Substance Name: Henicosafluoroundecanoic acid EC Number: 218 ...

Substance Name: Henicosafluoroundecanoic acid

EC Number: 218-165-4

CAS Number: 2058-94-8

MEMBER STATE COMMITTEE

SUPPORT DOCUMENT FOR IDENTIFICATION OF

HENICOSAFLUOROUNDECANOIC ACID

AS A SUBSTANCE OF VERY HIGH CONCERN BECAUSE OF ITS vPvB PROPERTIES

Adopted on 13 December 2012

SVHC SUPPORT DOCUMENT - HENICOSAFLUOROUNDECANOIC ACID

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CONTENTS

1 IDENTITY OF THE SUBSTANCE AND PHYSICAL AND CHEMICAL PROPERTIES ............................ 8

1.1 NAME AND OTHER IDENTIFIERS OF THE SUBSTANCE ..................................................................... 8 1.2 COMPOSITION OF THE SUBSTANCE ........................................................................................... 9 1.3 PHYSICO-CHEMICAL PROPERTIES............................................................................................. 9

2 HARMONISED CLASSIFICATION AND LABELLING ................................................................... 10

3 ENVIRONMENTAL FATE PROPERTIES...................................................................................... 10

3.1 DEGRADATION................................................................................................................... 10 3.1.1 Abiotic degradation ................................................................................................ 10 3.1.1.1 Hydrolysis .......................................................................................................... 10 3.1.1.2 Phototransformation/photolysis ......................................................................... 11 3.1.1.2.1 Phototransformation in air................................................................................... 11 3.1.1.2.2 Phototransformation in water .............................................................................. 12 3.1.1.2.3 Phototransformation in soil.................................................................................. 13

3.1.1.3 Summary and discussion on abiotic degradation................................................. 13 3.1.2 Biodegradation....................................................................................................... 14 3.1.2.1 Biodegradation in water ..................................................................................... 14 3.1.2.1.1 Estimated data .................................................................................................. 14 3.1.2.1.2 Screening tests ................................................................................................. 14 3.1.2.1.3 Simulation tests ................................................................................................ 15

3.1.2.2 Biodegradation in sediments............................................................................... 18 3.1.2.3 Biodegradation in soil ......................................................................................... 18 3.1.2.4 Summary and discussion on biodegradation ....................................................... 19

3.1.3 Summary and discussion on degradation ................................................................ 19 3.2 ENVIRONMENTAL DISTRIBUTION ........................................................................................... 20

3.2.1 Adsorption/desorption ........................................................................................... 20 3.2.2 Volatilisation .......................................................................................................... 20 3.2.3 Distribution modelling ............................................................................................ 20

3.3 BIOACCUMULATION ............................................................................................................ 20 3.3.1 Aquatic bioaccumulation......................................................................................... 21 3.3.1.1 Bioconcentration factor BCF................................................................................ 21 3.3.1.2 Biomagnification factors (BMFs) ......................................................................... 25 3.3.1.3 Trophic magnification factors (TMFs).................................................................. 30

3.3.2 Terrestrial bioaccumulation .................................................................................... 33 3.3.3 Summary and discussion of bioaccumulation .......................................................... 36

4 HUMAN HEALTH HAZARD ASSESSMENT .................................................................................. 37

5 ENVIRONMENTAL HAZARD ASSESSMENT................................................................................ 37

6 CONCLUSIONS ON THE SVHC PROPERTIES ............................................................................. 37

6.1 PBT, VPVB ASSESSMENT ..................................................................................................... 37 6.1.1 Assessment of PBT/vPvB properties – comparison with the criteria of Annex XIII.. 37 6.1.2 Bioaccumulation..................................................................................................... 39 6.1.3 Toxicity .................................................................................................................. 40 6.1.4 Summary and overall conclusions on the PBT, vPvB properties ............................... 40

6.2 CMR ASSESSMENT .............................................................................................................. 40 6.3 SUBSTANCES OF EQUIVALENT LEVEL OF CONCERN ASSESSMENT..................................................... 40

Annex I - Read-across approach ..................................................................................... 41

7 REFERENCES.......................................................................................................................... 48


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TABLES

Table 1: Substance identity ........................................................................................ 8 Table 2: Overview of physicochemical properties........................................................... 9 Table 3: Notified classification and labelling according to CLP criteria for C11-PFCA (query from October 2012) .......................................................................................................... 10 Table 4: Summary of photodegradation studies for C8-PFCA and its ammonium salt .......... 12 Table 5: Summary of screening tests for C12 PFCA, C14 PFCA, C8-9-PFCAs and the ammonium salt of C8-PFCA.......................................................................................................... 14 Table 6: Summary of simulations tests of C8-PFCA (PFOA) and its sodium and ammonium salt (APFO) ..................................................................................................................... 15 Table 7 measured Bioconcentration factors (BCF) of C11 –PFCA........................................ 23 Table 8 Calculated BCFs based on elimination rate and BMF values for C11-PFCA. The BMFs were measured by Martin et al. 2003b (please find details of the study in section 3.3.1.3). 23 Table 9 Bioaccumulation factors (BAF) of C11 –PFCA in different aquatic foodwebs detailed in section 3.3.1.2.......................................................................................................... 24 Table 10: Biomagnification factors (BMF) for C11-PFCA; if not indicated otherwise BMFs refer to whole body............................................................................................................... 28 Table 11: Trophic Magnification Factors (TMF) of C11-PFCA; if not indicated otherwise TMFs refer to whole body. .................................................................................................. 32 Table 12: TMF values for C11-C13-PFCA for a terrestrial food chain (Müller et al., 2011). ..... 34 Table 13: TMF values for C11-C13-PFCA for a terrestrial food chain (Müller et al., 2011). ..... 34 Table 14: CAS-Numbers and similarity of chemical structures of long chain PFCAs............. 41 Table 15: Basic substance information and physical chemical properties relevant to justify read across in the PBT assessment. .................................................................................... 43 Table 16 : Information on BCF, BMF and TMF of C9-14 PFCAs relevant to justify read across in the B assessment. ..................................................................................................... 46


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PREFACE

In four provided dossiers, the intrinsic properties of four perfluorinated carboxylic acids (PFCAs) are assessed: C11-14-PFCAs. Many studies are only available for structurally similar shorter chain PFCAs such as C8-PFCA and C9-PFCA. In those cases where studies on the particular substance are missing, studies from either shorter or longer chain PFCAs are used in the provided dossiers by applying read-across. Read-across is based on the structural similarities and on the physicochemical properties, which follow a regular pattern. All PFCAs contain a carboxylic acids group and a perfluorinated carbon chain. The only difference is the number of CF2-groups in this chain. Details on the read-across approach, i.e. showing the trend of physicochemical properties and the structural similarities are given in Annex I.


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Substance Name: Henicosafluoroundecanoic acid

EC Number: 218-165-4

CAS number: 2058-94-8

The substance is identified as a vPvB according to Article 57 (e).

Summary of how the substance meets the CMR (Cat 1A or 1B), PBT or vPvB criteria,

or is considered to be a substance giving rise to an equivalent level of concern

A weight of evidence determination according to the provisions of Annex XIII of REACH is used to identify the substance as vPvB. All available information (such as results of standard tests, monitoring and modelling, information from the application of the category and analog approach (grouping, read-across) and (Q)SAR results) was considered together in a weight of evidence approach. The individual results have been considered in the assessment with differing weights depending on their nature, adequacy and relevance. The available results are assembled together in a single weight of evidence determination.

Persistence:

Henicosafluoroundecanoic acid (C11-PFCA) has no degradation studies available.

Read across approach within C8-C14-PFCAs can be applied for the persistence assessment of these substances. C8-14-PFCAs contain a highly similar chemical structure, a perfluorinated carbon chain and a carboxylic acid group. The compounds differ only in the number of CF2-groups. As a result of comparing the experimental and estimated physico-chemical data of C8-PFCA (the analogue substance) with experimental and estimated data on C11-14-PFCAs it can be assumed that with increasing chain length water solubility decreases and the sorption potential increases (See Table 15 of the support document). It can be with a sufficient reliability stated that the behaviour of these chemicals follow a regular pattern.

Due to both structural similarity and a regular pattern of physico-chemical properties, C8-14-PFCAs may be considered as a group or a category of substances for the purpose of the PBT/vPvB assessment and the read-across approach can be applied within this group.

In general, the persistence of C11-C14-PFCAs can be explained by the shielding effect of the fluorine atoms, blocking, e.g., nucleophilic attacks to the carbon chain. High electronegativity, low polarizability and high bond energies make highly fluorinated alkanes to the most stable organic compounds. It is not expected that the carboxylic group in PFCAs alters this persistence of these chemicals. This fact is confirmed by a hydrolysis study which obtained a DT50 of >92 years for C8-PFCA in water. Screening studies of C8,9,12,14-PFCAs showed no biodegradation within 28 days. Non-standard abiotic degradation tests with C8-PFCA could not detect any degradation products under environmentally relevant conditions. Furthermore, screening biodegradation studies on C8,9,12,14-PFCAs and one non-standard anaerobic biodegradation simulation test with C8-PFCA provide evidence of high persistence. Additionally, elements of non-standard higher tier aerobic biodegradation studies on C8-PFCA provide further support that no biodegradation in water, soil and sediment occurs.

Therefore, based on the information summarized above, it is concluded that C11-PFCA is not degraded in the environment and thus fulfils the P- and vP- criteria in accordance with the criteria and provisions set out in Annex XIII of REACH.

Bioaccumulation

Results from bioconcentration and bioaccumulation studies in aquatic species:

The whole body BCF for C11-PFCA from the MITI flow-through study conducted with carp is in the range of 2300 – 3700 (National Institute of Technology and Evaluation, 2007). The study


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of Martin et al. considers that carcass BCFs closely approximate the whole-body BCF. This means the whole-body BCF would be approximately 2700. C11-PFCA would therefore be bioaccumulative but not very bioaccumulative if referred to whole body according to Annex XIII of REACH.

However, if taking into account that PFCAs show hepatotoxic effects, focus should rather be given to the higher accumulation potential in liver as bioaccumulation may be used as indicators for toxicity to organisms. Based on the findings there is a potential to reach critical levels that may elicit toxic effects over long-term exposures due to the high bioaccumulation potential to the liver.

BCF values of 5848-6326 (carcass) and 5246-5675 (liver) derived with two calculation methods from an experimental fish feeding study suggest that C11-PFCA is very bioaccumulative.

Other information on the bioaccumulation potential and ability of the substance to biomagnify

in the food chain:

Available field based bioaccumulation data show BAF values for C11-PFCA well above the trigger-value of 5000 for various species from different studies.

Various available field-BMFs and even TMFs for C11-PFCA provide evidence on that biomagnification of the substance takes place in nature between different trophic levels of food chains and from the bottom to the top of food chains.

Comparing the different field-BMF and TMF values for C9-14-PFCAs within single studies revealed a trend between chain length and BMF and TMF values, showing that C11-PFCA has a higher biomagnification potential than the longer chained homologues (see Data Matrix in Table 16 of the support document and Figures 1-4). The potential to magnify in the food web declined from C11-PFCA to C14-PFCA indicating that trophic magnification is more pronounced for C11-PFCA than for the longer chained PFCAs. Hence, although C11-PFCA seems to have in laboratory studies a lower bioaccumulation potential than its longer chained homologues, based on field data it has a higher potential to transfer through the food web as demonstrated by high BMF, TMF and BAF values. This observation especially points to the need of expert judgement based on weight-of-evidence.

A human biomonitoring study indicates that C11-PFCA is eliminated from human serum at a slower rate compared to its longer chained homologues C12-14-PFCAs. These results further support that C11-PFCA shows a higher bioaccumulation potential compared to C12-14-PFCAs in certain studies.

Conclusion on the bioaccumulation potential:

Based on only the BCFs derived from flow-through tests, C11-PFCA clearly fulfils the B criterion but some uncertainty remains whether it fulfils the vB criterion. BCFs derived from BMFs of a fish feeding study are above 5000. Although these laboratory data indicate that the vB criterion may be fulfilled, further confirmation is sought from field data.

Bioaccumulation factors derived in field studies are clearly above the trigger of 5000 which can be considered analogous to a BCF > 5000. Furthermore, the various available field-BMFs and TMFs for this substance provide evidence that biomagnification at high level takes place in nature. It is important to note that among C11-14-PFCAs, C11-PFCA based on field-BMFs and TMFs has the highest bioaccumulation potential indicating that trophic magnification is more pronounced for C11-PFCA than for the longer chained PFCAs. C12-14-PFCAs fulfill the vB criterion. Although for a large part of available information direct comparison with the vB criterion cannot be made, this information and especially the field data provide clear evidence on that the substance behaves in a way corresponding to bioaccumulation behaviour of substances which by direct comparison of the data to the vB criterion meet the criterion. Consequently, it is concluded that C11-PFCA fulfils both the B and the vB-criteria of REACH.


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Conclusion:

In conclusion, C11-PFCA is identified as a vPvB-substance according to Art. 57 (e) of REACH following the application of a weight of evidence determination using expert judgement by comparing all relevant and available information listed in Section 3 of Annex XIII of REACH with the criteria set out in Section 1 of the same Annex.

In conclusion, C11-PFCA is a vPvB-substance according to Article 57(e) of REACH.

The substance has not been registered under REACH yet.


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Justification

1 Identity of the substance and physical and chemical properties

1.1 Name and other identifiers of the substance

Table 1: Substance identity

EC number: 218-165-4

EC name: Henicosafluoroundecanoic acid

CAS number (in the EC inventory): 2058-94-8

CAS number: 2058-94-8

110961-38-1

CAS name: Undecanoic acid, 2,2,3,3,4,4,5,5,6,6,7,7,8,8,9,9,10,10,11,11,11-heneicosafluoro-

IUPAC name: Henicosafluoroundecanoic acid

Index number in Annex VI of the CLP Regulation

-

Molecular formula: C11HF21O2

Molecular weight range: 564.0909 g/mol

Synonyms: C11-PFCA

Perfluoroundecanoic acid Perfluoroundecylic acid

Structural formula:

CF3(CF2)9-COOH


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1.2 Composition of the substance

Name: Henicosafluoroundecanoic acid

Description: Mono-constituent substance

Degree of purity: Registration dossiers or other information on concentration ranges and on any impurities are not available.

1.3 Physico-chemical properties

The following physic-chemical properties are mainly obtained by calculation. No further information regarding the given values or other properties is available.

Table 2: Overview of physicochemical properties

Property Value Reference

Physical state at 20°C and 101.3 kPa

solid According to melting point

Melting/freezing point 112-114 °C Huang et al. 1987

Boiling point 238.4 °C at 101.325 kPa (calculated)

Kaiser et al. 2005

Vapour pressure 0.6 to 99.97 kPa (112 to 237.7°C) (calculated)

Kaiser et al. 2005

Water solubility 1.2E-4 g/L; pH 1 at 25°C

9.0E-4 g/L; pH 2 at 25°C

8.5E-3 g/L; pH 3 at 25°C

0.056 g/L; pH 4 at 25°C

0.14 g/L; pH 5 at 25°C

0.16 g/L; pH 6-10 at 25°C

Calculated using Advanced Chemistry Development (ACD/Labs) Software V11.02 (© 1994-2012 ACD/Labs)

Adsorption/desorption log Koc 3.19 – 3.41 Higgins and Luthy 2006

Dissociation constant 0.52±0.10 Calculated using Advanced Chemistry Development (ACD/Labs) Software V11.02 (© 1994-2012 ACD/Labs)

The above presented experimentally determined or calculated values are in good agreement with the physical chemical information, which is available for the homologues of PFCAs. With increasing chain length the melting and boiling point increase, while no significant change can be found for the vapour pressure and dissociation constant for the C11-14-PFCAs based on the calculations given in Table 3. The water solubility is predicted to decrease with increasing chain length. This is in agreement with the fact that the polarity of the substances decreases with an increasing chain length. It should be emphasised here that it is not possible to assess the calculated values of C11-14-PFCAs because there are factors like special conformation of the molecules which have an influence on the real values, but which have not been taken into account for the calculation.

However, the calculated values have dimensions which would be theoretically expected for the C11-14 PFCAs.


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2 Harmonised classification and labelling

C11-PFCA is not classified according to Annex VI of Regulation (EC) No 1272/2008 (CLP Regulation).

Twenty-eight notifications (3 aggregated notifications) have been submitted to the C&L Inventory.

Table 3: Notified classification and labelling according to CLP criteria for C11-PFCA (query from October 2012)

Classification Labelling

Hazard Class and Category Code(s)

Hazard Statement

Code(s)

Hazard Statement

Code(s)

Pictograms Signal Word Code(s)

Acute Tox. 4 Acute Tox. 4 Skin Irrit. 2 Eye Irrit. 2 Acute Tox. 4

H302 H312 H315 H319 H332

H302 H312 H315 H319 H332

GHS07 Wng

Not classified

Skin Irrit. 2 Eye Irrit. 2 STOT SE 3

H315 H319 H335

H315 H319 H335

GHS07 Wng

3 Environmental fate properties

3.1 Degradation

3.1.1 Abiotic degradation

3.1.1.1 Hydrolysis

There are no studies on the hydrolysis of C11-PFCA available. Based on the data given in Annex 1, results of studies of structurally similar substances of the same chemical group could be used to evaluate the hydrolysis of C11-PFCA.

Two studies are available on shorter chain lengths PFCAs. Hydrolysis of perfluorinated octanoic acid (C8-PFCA; PFOA) and its ammonium salt (APFO) (CAS-No: 335-67-1, 3825-26-1) and perfluorinated nonanoic acid (C9-PFCA; PFNA) (CAS No: 375-95-1) were analyzed. The studies are summarized in the following:

C8-PFCA is hydrolytically stable under relevant environmental conditions. One study has been discussed in the OECD SIDS Initial Assessment Report for C8-PFCA (PFOA), which has been copied here in italic letters (OECD, 2006):

The 3M Environmental Laboratory performed a study of the hydrolysis of APFO (3M Co.,

2001a) (Realiability = 1). The study procedures were based on USEPA’s OPPTS Guideline Document 835.2110; although the procedures do not fulfil all the requirements of the

guideline, they were more than adequate for these studies. Results were based on the

observed concentrations of APFO in buffered aqueous solutions as a function of time. The


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chosen analytical technique was high performance liquid chromatography with mass spectrometry detection (HPLC-MS).

During the study, samples were prepared and examined at six different pH levels from 1.5 to

11.0 over a period of 109 days. Experiments were performed at 50 °C and the results

extrapolated to 25 °C. Data from two of the pH levels (3.0 and 11) failed to meet the data

quality objective and were rejected. Also rejected were the data obtained for pH 1.5 because

ion pairing led to artificially low concentrations for all the incubation periods. The results for

the remaining pH levels (5.0, 7.0, and 9.0) indicated no clear dependence of the degradation

rate of PFOA on pH. From the data pooled over the three pH levels, it was estimated that the

hydrolytic half-life of PFOA at 25°C is greater than 92 years, with the most likely value of 235

years. From the mean value and precision of PFOA concentrations, it was estimated the

hydrolytic half-life of PFOA to be greater than 97 years.

A newer study showed no decomposition of C8-9-PFCAs in hot water in absence of S2O82-. After

the addition of S2O82- to the reaction system efficient decomposition of PFCAs has been

observed at 80 °C. After a reaction time of 6 hours, C8-PFCA and C9-PFCA were decomposed completely. The reaction products were mainly F- and CO2 at a yield of 77.5 % ((moles of F

- formed)/(moles of fluorine content in initial PFOA)) and 70.2 % ((moles of CO2 formed)/(moles of carbon content in initial PFOA)), respectively for C8-PFCA. For C9-PFCA the reaction products were mainly F- and CO2 at a yield of 88.9 % and 75.2 %, respectively. Short chain PFCAs were a minor reaction product. However, at higher temperatures (150°C) 12.3% of the initial C8-PFCA remained and the yields of F- and CO2 were 24.6 and 37.0 %, respectively (Hori et al., 2008) (Reliabilty = 2).

The water solubility of C11-PFCA is lower than those of C8,9-PFCA, which can be explained by the expanded fluorinated carbon chain (Annex 1). However, the stability of the PFCAs is mainly based on the stability of the highly fluorinated carbon chain (Siegemund et al., 2000). Since C8-PFCA is hydrolytically stable, we estimate a comparable hydrolytically stability also for C11-PFCA.

Based on the read across rationale described in Annex 1, data on C8-PFCA is used as evidence for C11-PFCA to conclude that it is hydrolytically stable under environmental conditions.

3.1.1.2 Phototransformation/photolysis

Direct photolysis of a carbon fluorine chain is expected to be very slow, with stability expected to be sustained for more than 1000 years (Environment Canada, 2012).

3.1.1.2.1 Phototransformation in air

There are no studies on phototransformation for C11-PFCAs in air available. However, studies on C8-PFCA exist and are summarized below:

The following information was copied from the OECD SIDS Initial Assessment Report for C8-PFCA (PFOA) (OECD, 2006):

Hurley et al. determined the rate constants of the reactions of OH radicals with a homologous

series of perfluorinated acids (from trifluoroacetic acid to nonafluoropentanoic acid) in 700 Torr

of air at 296 K (Hurley et al., 2004). For C3 to C5 chain length had no discernible impact on the

reactivity of the molecule. The rate constant k(OH + F(CF2)nCOOH) = (1.69±0.22)×10-

13 cm3 molecule-1 s-1 for n = 2, 3, 4, respectively. Atmospheric lifetimes of F(CF2)nCOOH with

respect to reaction with OH radicals are estimated to be approximately 230 days for n = 1 and

130 days for n > 1. (Calculation of lifetime by comparison with CH3CCl3 (half-life 5.99 years,

k = 1.0×10-14 cm3 molecule-1 s-1). The authors conclude, that the major atmospheric loss

mechanism of perfluorinated carboxylic acids is dry and wet (particle mediated) deposition

which occur on a time scale which is probably of the order of 10 days. Reaction with OH is a

minor atmospheric loss mechanism for perfluorinated carboxylic acids


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3.1.1.2.2 Phototransformation in water

The photochemical decomposition of long-chain PFCAs in water by use of persulfate ion (S2O82-

) in water (C9-PFCA) and in an aqueous/liquid CO2 biphasic system (C9-11-PFCAs) was examined by Hori et al. (Hori et al., 2005b) (Reliability = 2). In water and in the absence of S2O8

2- (direct photolysis) C9-PFCA decomposition of 64.5 % was determined. In the presence of

S2O82- the decomposition increased to 100%. The decompositions after 12 hours in the biphasic

system were 100% for C9-PFCA and C10-PFCA, and 77.1% for C11-PFCA. The reaction product was mainly F- (66.2 %, 73.4 % and 46.35 % of (moles of F- formed)/(moles of fluorine content in initial PFCA)) and the minor reaction products were short-chain PFCAs. Since the conditions in this study are not environmentally relevant, we did not describe this study in detail.

In addition to the study of Hori et al, further studies are available for C8-PFCA (PFOA) and its ammonium salt APFO (see table 4).

Table 4: Summary of photodegradation studies for C8-PFCA and its ammonium salt

Test

Substance

Result Remarks Reliability Reference

Ammonium salt of C8-PFCA

No photodegradation Direct photolysis 2 (OECD, 2006);(3M Co.,

1979)

No photodegradation Direct and indirect (H2O2; synthethic humic water, Fe2O3) photolysis


Estimated half-life > 349 days

Indirect photolysis (Fe2O3)

1 (OECD, 2006);(3M Co.,

2001b)

Short wave length (<300 nm) used for irradiation → limited relevance for an aqueous environment

44.9% of the initial PFOA was decomposed after 24 hours

Direct photolysis;

0.48 MPa O2


Indirect photolysis (H2O2); 0.48 MPa O2

C8-PFCA

100% of the initial PFOA was decomposed after 24 hours

Indirect photolysis (tungstic heteropolyacid photocatalyst); 0.48 MPa O2

2 (Hori et al., 2004)

Short wave length (<300 nm) used for irradiation → limited relevance for an aqueous environment


Direct photolysis; 0.48 MPa O2

C8-PFCA

100% of the initial PFOA was decomposed after 4 hours

Indirect photolysis (S2O82-); 0.48

MPa O2

2 (Hori et al., 2005a)

The following information was copied from the OECD SIDS Initial Assessment Report for C8-PFCA (PFOA; APFO is the ammonium salt of C8-PFCA) (OECD, 2006):

Direct photolysis of APFO was examined in two separate studies (3M Co., 1979; 3M Co., 2001b) and photodegradation was not observed in either study. In the 3M (1979) study, a

solution of 50 mg/l APFO in 2.8 litres of distilled water was exposed to simulated sunlight at


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22±2 ºC. Spectral energy was characterized from 290-600 nm with a max output at ~360 nm. Direct photolysis of the test substance was not detected.

In the 3M (3M Co., 2001b) study, both direct and indirect photolysis was examined utilizing

techniques based on USEPA and OECD guidance documents. To determine the potential for

direct photolysis, APFO was dissolved in pH 7 buffered water and exposed to simulated

sunlight. For indirect photolysis, APFO was dissolved in 3 separate matrices and exposed to

simulated sunlight for periods of time from 69.5 to 164 hours. These exposures tested how

each matrix would affect the photodegradation of APFO. One matrix was a pH 7 buffered

aqueous solution containing H2O2 as a well-characterized source of OH radicals. This tested the

propensity of APFO to undergo indirect photolysis. The second matrix contained Fe2O3 in water

that has been shown to generate hydroxyl radicals via a Fenton-type reaction in the presence

of natural and artificial sunlight. The third matrix contained a standard solution of humic

material. Neither direct nor indirect photolysis of APFO was observed based on loss of starting

material. Predicted degradation products were not detected above their limits of quantitation.

There was no conclusive evidence of direct or indirect photolysis whose rates of degradation

are highly dependent on the experimental conditions. Using the iron oxide (Fe2O3)

photoinitiator matrix model, the APFO half-life was estimated to be greater than 349 days.

According to Hori et al., aqueous solutions of PFOA absorb light strongly from the deep UV-

region to 220 nm (Hori et al., 2004). A weak, broad absorption band reaches from 220 to

270 nm (no absorption coefficient stated). The irradiation of a 1.35 mM PFOA solution (29.6

µmol) in water (under 0.48 MPa of oxygen) with light from a xenon-mercury lamp (no radiant

flux stated) for 24 hours resulted in a ca. 44.9 % reduction (13.3 µmol) of PFOA concentration.

Concentrations of CO2 and fluoride increased simultaneously. Small amounts (0.1-5 µmol) of

short chain perfluorinated hydrocarbon acids (C2-C7) were detected. The addition of the

photocatalyst tungsten heteropolyacid ([PW12O40]-) or persulfate (S2O8

2-) (Hori et al., 2005a) accelerates the reaction rate. Due to the short wave length used for irradiation (< 300 nm) the

photodegradation described may be of limited relevance for an aqueous environment but may be used as a technical process.

3.1.1.2.3 Phototransformation in soil

No data available

3.1.1.3 Summary and discussion on abiotic degradation

In general, the perfluorinated carboxylic acids are very stable. Since there are no degradation studies (under relevant environmental conditions) on the C11-PFCA available, data from similar substances need to be considered and discussed. Based on the data given in Annex 1, results of studies of structurally similar substances of the same chemical group could be used to evaluate the abiotic degradation of C11-PFCA.

The data on C8-PFCA indicate that abiotic degradation in the atmosphere is expected to be slow (atmospheric lifetime = 130 days; conclusion by analogy from short-chain perfluorinated acids). Under relevant environmental conditions C8-PFCA is hydrolytically stable (DT50 > 92 years) and do not undergo direct photodegradation in natural waters. The estimated DT50 for indirect photolysis is 349 days.

Based on the read across rationale described in Annex 1, data on C8-PFCA are used as evidence for C11-PFCA to conclude that it is stable under environmental conditions and abiotic degradation is expected to be as low as for the chemically similar substance C8-PFCA.


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3.1.2 Biodegradation

3.1.2.1 Biodegradation in water

3.1.2.1.1 Estimated data

For the C9-PFCA a half-life in water of 2477 days and a half-life in soil of 4954 days was estimated (Lambert et al. 2011).

3.1.2.1.2 Screening tests

There are no studies available for the C11 PFCA. Based on the data given in Annex 1, results of studies of structurally similar substances of the same chemical group could be used to evaluate the hydrolysis of C11-PFCA.

One study is available for the structurally similar C12,14-PFCAs, respectively. Moreover, some studies are available on C8,9-PFCAs and the ammonium salt of C8-PFCA. The results are summarized in Table 5.

Table 5: Summary of screening tests for C12 PFCA, C14 PFCA, C8-9-PFCAs and the ammonium salt of C8-PFCA

Test

substance

Method Result Reliability Reference

C8-PFCA OECD 301 C 5 % in 28 days 2 (National Institute of Technology and Evaluation, 2007)


OECD 301 C 7 % in 28 days 2 (National Institute of Technology and Evaluation, 2007)

C12-PFCA OECD 301 C No biodegradation in 28 days

2 (National Institute of Technology and Evaluation, 2002)

C14-PFCA OECD 301 C No biodegradation in 28 days

2 (National Institute of Technology and Evaluation, 2002)


OECD 301 B 13 % in 28 days 2 (OECD, 2007), (DuPont Co., 1997)

C8-PFCA OECD 301 F No biodegradation in 28 days

2 (Stasinakis et al., 2008)

C9-PFCA OECD 301 F No biodegradation in 28 days

2 (Stasinakis et al., 2008)


Shake culture test modelled after the Soap and Detergent Association´s presumptive test for degradation

No biodegradation after 2.5 months

2 (OECD, 2006), (3M Co., 1978)


15

A number of studies for the C8-PFCA (PFOA) and its ammonium salt APFO were already discussed in the OECD SIDS Assessment Report (OECD, 2006). The following text in italic letters was copied from there:

Using an acclimated sludge inoculum, the biodegradation of APFO was investigated using a

shake culture study modeled after the Soap and Detergent Association's presumptive test for

degradation (3M Co., 1978). Both thin-layer and liquid chromatography did not detect the

presence of any metabolic products over the course of 2 1/2 months indicating that PFOA does

not readily undergo biodegradation. In a related study, 2.645 mg/l APFO was not measurably

degraded in activated sludge inoculum (Pace Analytical, 2001). Test flasks were prepared

using a mineral salts medium, 1 ml methanol, and 50 ml settled sludge. Analysis was

conducted with a HPLC/MSD system. Although the results were deemed unreliable due to a

lack of description of experimental protocols or indications of a high degree of experimental

error, several other studies conducted between 1977-1987 also did not observe APFO

biodegradation (Pace Analytical, 1987; 3M Co., 1985; 3M Co., 1980; 3M Co., 1979). In

addition, a study conducted by Oakes et al.) . In addition, a study conducted by Oakes et al.

indicated little biotic or abiotic degradation of PFOA on a time scale of 35 days, i.e., the PFOA

exposure concentrations were stable over time and ranged from 84.5 % to 114.5 % of the

initial concentrations (Oakes et al., 2004).

In a 28 day ready biodegradability test (OECD 301 C) using 100 mg/L C12-PFCA, C14-PFCA, C8-PFCA and its ammonium salt, respectively, and 30 mg/L activated sludge non-biodegradability was demonstrated. Only 5 % (C8-PFCA) and 7% (ammonium salt of C8-PFCA) degradation was observed by BOD. For C12-PFCA and C14-PFCA no biodegradation was observed (National Institute of Technology and Evaluation, 2007).

In a further test of ready biodegradability (OECD 301 F) biodegradation of neither C8-PFCA nor C9-PFCA was observed in 28 days (Stasinakis et al., 2008).

In summary, on the basis of the available screening tests, C8,9,12,14 PFCAs are not readily biodegradable. Based on the read across rationale described in Annex 1, data on C8,9,12,14-PFCAs are used as evidence for C11-PFCA to conclude that it is not readily biodegradable.

3.1.2.1.3 Simulation tests

For C11-PFCAs no experimental degradation test is available. Based on the data given in Annex 1, results of studies of structurally similar substances of the same chemical group could be used to evaluate the hydrolysis of C11-PFCA.

Therefore, test results for C8-PFCA are discussed shortly in the following.

No environmental half-lives for C8-PFCA have been reported, even in the cases where corresponding tests have been performed (see table 6).

Table 6: Summary of simulations tests of C8-PFCA (PFOA) and its sodium and ammonium salt (APFO)

Test

substance

Method Result Reliability Reference

C8-PFCA Closed-loop systems in laboratory scale;

Aerobic and anaerobic conditions

No elimination 3 (Meesters and Schroeder, 2004; Schröder, 2003)


Biodegradation in mixed bacterial culture and activated sludge

Aerobic conditions

< 0.6 % of 14CO2 was detected after 28 days

4 (Wang et al., 2005)


16

Sodium salt of C8-PFCA

Microcosm study

Aerobic conditions

No significant dissipation from water column after 35 days (initial concentration 0.3 mg/L; 1mg/L; 30 mg/L)

32% dissipation in 35 days (initial concentration 100 mg/L)

3 (Hanson et al., 2005)

C8-PFCA/ ammonium salt of C8-PFCA

1.Preliminary screening:

C8-PFCA serves as an electron acceptor under anaerobic conditions (in combination with different inocula)

2. Hypothesis refinement:

14C C8-PFCA serves as an electron acceptor under anaerobic conditions

No significant consumption of the initial C8-PFCA during 110 – 259 days

No loss of ammonium salt of C8-PFCA No production of 14CO2 No detection of radiolabel transformation products

2 (Liou et al., 2010)

In the OECD SIDS Initial Assessment Report it was concluded that C8-PFCA (PFOA) is not expected to undergo biodegradation (OECD, 2006). The following text in italic letters was copied from there:

Schroeder (2003), and Meesters and Schroeder (2004) investigated the biochemical

degradation of PFOA in sewage sludge in laboratory scale reactors. After 25 days under aerobic

conditions PFOA (initial concentration 5 mg/l) was not eliminated by metabolic processes,

mineralization processes or by adsorption (Meesters and Schroeder, 2004; Schröder, 2003).

Wang et al. studied the biodegradation of fluorotelomer alcohols. However, 14C-labelled C8-PFCA ammonium salt was used as starting material in this study, too. The authors analyzed the headspace of sealed vessels containing mixed bacterial cultures and vessels containing activated sludge from a domestic sewage treatment plant under continuous air flow. The mixed bacterial culture from industrial wastewater treatment sludge was enriched using 8:2 telomere alcohol and 14C-labelled C8-PFCA ammonium salt, respectively. However, for using C8-PFCA ammonium salt as a starting material no detailed information are available from the report. The authors describe that potential biodegradation products were separated and quantified by LC/ARC (on-line liquid chromatography/accurate radioisotope counting). Transformation products were identified by quadrupole time of flight mass spectrometry. Only <0.6 % of 14CO2 was detected after 28 days. The report contains no graphs or further data to re-evaluate this statement. Although the study seems to be very well documented for 14C labelled 8:2 FTOH, we can only flag the study with a reliability of 4, since details on C8-PFCA ammonium salt are not available. The documentation for the results obtained with C8-PFCA is missing in the report. However the result indicates that C8-PFCA ammonium salt is not biodegradable within 28 days (Wang et al., 2005).

Hanson et al. performed a microcosm study. Microcosms were approximately 1.2 m deep with a water depth of 1 m, a diameter of 3.9 m, and a surface area of 11.95 m². Each microcosm had a capacity of approximately 12m³ of water. Sediment consisted of a 1:1:1 mixture of sand, loam and organic matter (mainly composted manure). The total carbon content of the sediment was 16.3%. Microcosms were circulated for 2 weeks from a well-fed irrigation pond prior to the experiments. Nominal concentrations of 0.3, 1, 30, and 100 mg/L C8-PFCA, as the sodium salt, plus controls were added to the microcosms. Each exposure was randomly assigned to three separate microcosms from a total of 15 microcosms. Immediately prior to treatment, water flow into each microcosm from the main irrigation pond ceased, creating a closed system relative to the other microcosms and the irrigation pond.


17

Water chemistry and PFOA analysis were taken at the same time on a regularly basis. Temperature and dissolved oxygen content were measured daily. Water samples were collected with a metal integrated water column sampler. Integrated subsamples from at least 4 randomly selected locations in each microcosm were collected to a total volume of 4 L. Samples were stored at 4 °C until analysis. Water samples were analyzed by ion chromatography. The mobile phase was 0.5 mM NaOH, 5 % methanol, and 5% acetonitrile with a flow rate of 0.4 mL/min. Injection volumes varied from 5, 10, 75, and 200 µl for the 100, 30, 1 and 0.3 mg/L microcosms, respectively. For each set of samples analyzed five standards and one quality control sample were included at the beginning of each run and again at the end. Radioactive labelling was not performed. Over a 35-day field study C8-PFCA showed no significant dissipation from the water column. However, at the highest concentration (100 mg/L) a partitioning from the water column into other compartments is suspected (32% dissipation in 35 days) (Hanson et al., 2005). Since the documentation of the procedure was insufficient in our opinion the study is not reliable (reliability 3).

Liou et al. investigated the anaerobic biodegradability of C8-PFCA respectively its ammonium salt. In a two-phase experiment (preliminary screening, hypothesis refinement) the use of C8-PFCA as a physiological electron acceptor (electron donator: acetate, lactate, ethanol or hydrogen gas) was studied. Additionally, the possibility of co-metabolism of C8-PFCA during reductive dechlorination of trichloroethene and during various physiological conditions (aerobic, nitrate-reducing, iron-reducing, sulfate reducing, and methanogenic) was analyzed. Five different inoculums were used (from a municipal waste-water treatment plant, industrial site sediment, an agricultural soil, and soils from two fire training areas). Environmental samples used as inoculum sources in the biodegradation experiments were aseptically gathered (sterile spatula) placed in 0.5 L sterilized canning jars (filled to the brim), stored on ice in the field, and maintained at 4 °C before being transferred to an anaerobic hood where samples were degassed and dispensed as slurries in biodegradation assays. Soils and sludges were gathered from: the Ithaca sewage treatment plant; a water-saturated drainage ditch adjacent to the DuPont Chambers Works waste treatment facility in Salem County, New Jersey, previously shown to carry out reductive dechlorination (Fung et al., 2009); the Cornell agricultural field station (Collamer silt loam, Ithaca, NY), the Ithaca fire training facility, and the Rochester, NY fire training facility (the latter two sites were chosen due to potential contamination with fluorinated fire retardant chemicals) (Liou et al., 2010).

For the serum bottle –based biodegradation assays treatments occurred in triplicats (160 ml serum bottles with 100 mL of media; live ± C8-PFCA and abiotic controls, autoclaved for 1 h). For the 14C-PFOA experiments, 15-mL serum bottles were utilized (50% O2-free N2 headspace, 50% inoculated anaerobic test medium) with non-radioactive C8-PFCA and

14C- C8-PFCA (4.5 lCi/mL test medium) to give a final concentration of 100 mg/L C8-PFCA medium. For establishing the various terminal electron-accepting processes, a standard anaerobic procedure was used. The anaerobic mineral salts buffer (plus vitamins and trace minerals) was used as diluents for the various inoculums sources (5% wt/volume) with addition of electron donors (10 mM sodium acetate ± 40 mM sodium lactate or 0.6 mM ethanol or 2 atm H2) or electron acceptors [O2 as air headspace or O2- free N2 headspace in each serum bottle with additions of 30 mM nitrate or 4 mg/ mL_FeOOH or 10 mM sulfate or 0.4 mM trichloroethene (TCE) or no addition (for the methanogenic treatment)]. Samples (1.0 mL) were periodically removed from each serum bottle, placed in 4-mL glass vials sealed with Al-backed caps, immediately mixed with an equal volume of methanol and stored at _20 °C until analyzed. Accumulated batches of samples from serum vials were analyzed for concentrations of PFOA, 14C- C8-PFCA, fluoride, nitrate, sulfate, and potential C8-PFCA transformation products. Headspace gases were sampled with a gas-tight syringe (250 mL) and analyzed for TCE, vinyl chloride and methane. In the radiotracer study, dissolved 14C activity in the anaerobic medium and in the 0.4 N KOH solution retrieved from the internal reservoir to trap 14CO2 were determined by scintillation counting. To assay potential microbial inhibition by C8-PFCA, triplicate serum- bottle assays inoculated with 5% Ithaca sewage were prepared, as above. Anaerobic preparations (±100 ppm C8-PFCA) were assayed for methanogenesis. Aerobic preparations containing 15 ppm naphthalene were sampled as above and analyzed by high-performance liquid chromatography (HPLC). After filtration through nylon acrodisc filters, naphthalene was separated at room temperature. Methanol–water (1:1) was the mobile phase at a flow rate of 1.5 mL/ min. The


18

eluent was monitored by UV VIS at 340 nm. Quantification was done by comparison to authentic standards) (Liou et al., 2010). C8-PFCA quantification was performed by LC/MS/MS following a standard procedure. Potential C8-PFCA metabolites were screened by applying LC/MS.

In no combination of the inoculum source, electron donator or physiological conditions a significant percentage of the initial C8-PFCA (100 ppm and 100 ppb) was consumed (110 - 259 days). In a test with 14C labelled C8-PFCA ammonium salt (inoculum = sewage), no loss of C8-PFCA ammonium salt was detected, no 14CO2 was produced and no radiolabelled C8-PFCA ammonium salt transformation product was indicated. Co-metabolism of C8-PFCA during reductive dechlorination of trichlorethene was suggested by a drop in C8-PFCA concentration in the 100 ppb treatment after a 65-d incubation. However, extensive analysis failed to determine corroborating transformation products. In summary, under conditions which were examined in this study, C8-PFCA is environmentally persistent (Liou et al., 2010).

In conclusion, the one non-standard aerobic degradation simulation study and one non-standard anaerobic degradation simulation study on C8-PFCA demonstrate the high persistence of the compound. Based on the read across rationale described in Annex 1, data on C8-PFCAs can be used as evidence of persistence for C11-PFCA.

3.1.2.2 Biodegradation in sediments

No data available

3.1.2.3 Biodegradation in soil

There are no degradation studies on C11-PFCA available. Based on the data given in Annex 1, results of studies of structurally similar substances of the same chemical group could be used to evaluate the hydrolysis of C11-PFCA.

A number of studies are available for C8-PFCA (PFOA) which was already discussed in the OECD SIDS Initial Assessment Report. The following text was copied from there (italic letters) (OECD, 2006):

Moody and Field (1999) conducted sampling and analysis of samples taken from groundwater

1 to 3 meters below the soil surface in close proximity to two fire-training areas with a history

of aqueous film forming foam use. Perfluorooctanoate was detected at maximum

concentrations ranging from 116 to 6750 µg/l at the two sites many years after its use at

those sites had been discontinued. These results suggest that PFOA can leach to groundwater

(Moody and Field, 1999).

Extensive site specific monitoring of soil and ground water concentrations of PFOA and related

substances was conducted by 3M, DuPont Daikin and others. PFOA in soil has been shown to

persist for decades and to be a long term source of groundwater and surface water

contamination (see for example (DuPont Co., 2003; 3M Co., 2005)).

At the DuPont Washington Works site soil contaminated by perfluorochemical waste has been shown to contain ppm levels of PFOA 3 decades after application ceased. The underlying

groundwater also contains ppm levels of PFOA (DuPont Co., 1999).

Extensive field monitoring data generated by 3M at the Decatur, AL site have also shown that

PFOA is persistent in soils. Soil samples were collected from a former sludge application area of

the 3M Decatur, AL facility also show soil contamination and underlying groundwater

contamination up to ppm levels decades after application ceased.

Moody et al. investigated groundwater at a former fire-training area at Wurtsmith Air Force Base which was used between 1950s and 1993. Before sampling, the soil and groundwater in the area has been studied in detail. Groundwater samples were collected from two types of


19

monitoring wells. All samples were collected in high density polypropylene bottles. Samples were shipped on ice without preservation and stored at 4 °C prior to analysis. Perfluorocarboxylate concentrations were measured as described in the following: Strong anion exchange disks were used to extract perfluorocarboxylates (6 to 8 carbons) from groundwater. The perfluorocarboxylates were simultaneously eluted from the disks and derivatized to their methyl esters by treatment with iodomethane for direct analysis by electron impact gas chromatography-mass spectrometry (GC-MS). A single analysis was conducted for each groundwater sample. The detection limit (defined as a signal-to-noise ratio greater than 3) and quantification limit (defined as a signal-to-noise ratio greater than 10) for perfluorocarboxylates were 3 and 13 mg/ L, respectively, using 2-chlorolepidine as the internal standard. Additionally, electron capture negative chemical ionization GC-MS was employed to confirm the identity of PFOA, in groundwater samples (Moody et al., 2003). Depending on the location of sampling, the concentrations of C8-PFCA were between 8 and 105 µg/L in groundwater. The authors estimated that perfluorinated surfactants had been in the groundwater for at least five years and possibly for as long as 15 years. This showed that degradation of C8-PFCA was negligible under the environmental conditions at this site (for both soil and groundwater) (Reliability = 2) (Moody et al., 2003).

The anaerobic biodegradability of C8-PFCA and its ammonium salt, respectively, in soil from two fire training areas was investigated by Liou et al. (see above 3.1.2.1.3 Simulation tests). No significant amount of the initial PFOA was dissipated after 259 days.

In conclusion, the available data on C8-PFCA demonstrate the high persistence of the compound. Based on the read across rationale described in Annex 1, data on C8-PFCAs can be used as evidence of persistence for C11-PFCA.

3.1.2.4 Summary and discussion on biodegradation

Screening studies for C11-PFCA are not available. However, results from screening studies of C8,9,12,14-PFCAs used for read across approach as described in Annex1 indicate that structurally similar compounds are not readily biodegradable. The results of one non-standard aerobic biodegradation simulation test, one non-standard anaerobic biodegradation simulation test and field monitoring data on C8-PFCA from contaminated sites provide evidence that no biodegradation in water, soil and sediment occurs. Since the stability of PFCAs is in general mainly based on the stability of the fluorinated carbon chain it can be concluded that also for C11-PFCA no biodegradation in water, soil and sediment can be expected. Thus, it can be assumed that C11-PFCA is persistent as well.

3.1.3 Summary and discussion on degradation

For C11-PFCA no experimental data on degradation are available. Therefore, data from chemically similar compounds should be considered in a read-across approach (please see Annex 1 for further details). The degradation potential of substances differing only in the number of carbons in the fluorinated carbon chain has been analyzed in some studies. Generally, it is known that the bond between carbon and fluorine is one of the most stable ones in organic chemistry.

A number of studies for the shorter chain C8-PFCA show that this substance is very persistent and does not undergo abiotic or biotic degradation at all under relevant environmental conditions.

Abiotic degradation

The data on C8-PFCA indicate that abiotic degradation in the atmosphere is expected to be slow (atmospheric lifetime = 130 days). The hydrolytic half-life of C8-PFCA at 25°C is greater than 92 years, with the most likely value of 235 years under relevant environmental conditions (3M Co., 2001a). No photodegradation of C8-PFCA has been observed in studies conducted


20

under relevant environmental conditions. The estimated DT50 for indirect photolysis is 349 days.

Biotic degradation

Standard screening tests are available for C8,9,12,14-PFCAs. No biodegradation at all has been observed for C9,12,14-PFCAs within 28 days. For C8-PFCA test results differ from “no biodegradation to 13% biodegradation of the ammonium salt. Thus, it can be concluded that C8,9,12,14-PFCAs are not readily biodegradable.

For C8-PFCA a non-standard aerobic biodegradation simulation test, one non-standard anaerobic biodegradation simulation test and field monitoring data from contaminated sites provide evidence that no biodegradation in water, soil and sediment occurs.

Conclusion

PFCAs are synthetic compounds which contain a structural feature: a perfluorinated carbon chain combined with a carboxylic group. The chemical structure of these compounds differs only in the number of perfluorinated carbons in the carbon chain.

The stability of organic fluorine compounds has been described in detail by Siegemund et al., 2000: When all valences of a carbon chain are satisfied by fluorine, the zig-zag-shaped carbon skeleton is twisted out of its plane in the form of a helix. This situation allows the electronegative fluorine substituents to envelope the carbon skeleton completely and shield it from chemical attack. Several other properties of the carbon-fluorine bond contribute to the fact that highly fluorinated alkanes are the most stable organic compounds. These include polarizability and high bond energies, which increase with increasing substitution by fluorine. The influence of fluorine is greatest in highly fluorinated and perfluorinated compounds. Properties that are exploited commercially include high thermal and chemical stability (Siegemund et al., 2000).

Based on their molecular properties it is, thus, not surprising, that researchers could not measure degradation of the intensively studied C8-PFCA or its salts. Considering the organic chemistry of this substance group it seems to be very likely that a carbon chain being some CF2-groups longer is as persistent as a shorter chain C8-PFCAs. We therefore conclude that C9-14-PFCAs are as resistant to degradation as it has been shown for C8-PFCA.

In summary, using the described read-across approach, we conclude that C11-PFCA is a very persistent synthetic compound which is resistant to abiotic and biotic degradation.

3.2 Environmental distribution

3.2.1 Adsorption/desorption

Not relevant for the SVHC identification of the substance in accordance with Article 57 (e).

3.2.2 Volatilisation


3.2.3 Distribution modelling



21

3.3 Bioaccumulation

3.3.1 Aquatic bioaccumulation

3.3.1.1 Bioconcentration factor BCF

Bioconcentration is the process by which a chemical is accumulated by an organism as a result of exposure to the chemical in water – it often refers to a condition usually achieved under laboratory and steady state conditions. The BCF is typically calculated as the ratio of the measured concentrations of the chemical in the organism and the water once a steady state has been achieved

Water

Biota

c

cBCF =

The BFC can alternatively be determined kinetically by using the uptake rate k1 and the depuration rate k2:

2

1

k

kBCF =

There are two studies available which determined the BCFs of C11-PFCA.

In the first study carp were exposed to C11-PFCA (National Institute of Technology and Evaluation, 2007). The test was conducted in accordance with the OECD 305 guideline. The test was conducted in a flow through test system, the concentration of the test substance were 1 µg/L and 0.1µg/L and below the predicted water solubility value. The concentrations were analytically checked via HPLC-MS. The pH was within the range 7.8 to 8.1. The uptake period was 60 days. The depuration period was 45 days. Steady state BCFs were in the range from 2300 - 3700 for C11-PFCA (Table 7). In the second study rainbow trout were exposed in a flow-through system for 12 days followed by a depuration time of 33 days in fresh water to determine tissue distribution and bioconcentration (Martin et al., 2003a). For determination of bioconcentration, juvenile fish (5-10g) were exposed simultaneously to PFCAs of varying chain length. No adverse effects were observable based on fish mortality, growth and liver somatic index. The exposure concentration of each PFCA was analytically checked. PFCA concentrations were stable throughout the uptake phase. For C11-PFCA the measured concentration was 0.48 µg/L with a relative standard deviation of 26%. There was an initial decrease between 0.25 h and 24 h which is considered to be caused by the rapid uptake of the PFCAs. At 7 occasions during uptake period and 9 occasions during depuration phase, three fish from the exposure tank and one fish from the control were removed to determine the kinetics of uptake and depuration. The BCFs (carcass, blood and liver) were determined on the basis of the uptake and depuration kinetics and results are given in Table 7. All tissue concentrations were corrected for growth dilution. Additionally, for the tissue distribution study, four immature trout (30 – 48 g) were exposed in separate tanks but under the same uptake conditions (Martin et al., 2003a).

This tissue distribution study showed that unlike lipophilic organic compounds PFCAs did not preferentially accumulate in adipose tissue. Hence a lipid-normalisation of the BCFs would not be reasonable. C11-C14-PFCA concentrations were highest in blood, kidney, liver and gall bladder and low in the gonads, adipose and muscle tissue. Within the blood, the plasma contained between 94 – 99 % of PFCA, with only a minor fraction detectable in the cellular fraction. Recovery from hearts and spleen was low (<10%). Based on high blood, liver and gall bladder concentrations and slow depuration the authors assume that PFCA enter the


22

enterohepatic recirculation in fish. That means the compounds are continuously transferred between the different organs (Martin et al., 2003a).

Additionally, BCFs were calculated for different body compartments. Though, bioaccumulation should preferably be based on whole body. According to the authors carcass BCFs closely approximate the whole-body BCF. However, compartment-specific BCFs can be more relevant if there is a potential for direct organ-specific toxicity. PFCAs cause hepatomegaly in rodents (Kudo et al., 2000) which is an indicator for hepatotoxicity. This study investigated PFCAs with 7–10 carbon chain lengths. PFCAs having a longer perfluorinated carbon chain showed a more potent induction of hepatomegaly. C11-PFCA was not investigated. However due to its structural similarity, with an additional fluorinated carbon, it can be assumed that also C11-PFCA may be hepatotoxic. Thus, from a toxicological perspective, BCFs based on concentrations in individual organs, such as the liver, may be more relevant when the potential for direct organ-specific toxicity (i.e., liver toxicity) is being predicted. No statistically significant difference was found between the liver somatic index of exposed and control fish. However, bioaccumulation tests are not designed for showing toxic effects. BCF values may be estimated using biomagnification data. Based on the assumption the elimination kinetics are the same regardless of the uptake pathway, a kinetic BCF can be calculated. For this the measured elimination-rate from a fish feeding study and a modelled uptake rate for uptake through the gills is used. There are various mechanistic models available to roughly estimate the uptake rate. An overview is given in Crookes and Brooks (2011). These models depend on physical-chemical input parameters such as the log KOW. As this parameter cannot be sufficiently estimated for PFCAs these approaches should not be used (Weisbrod et al., 2009). However, based on a study with in vivo and isolated perfused gills Sijm et al. suggested an approach, which is independent of the log KOW and describes an uptake rate dependent on fish weight only (Sijm et al. 1995). This approach has been suggested as potentially suitable for the estimation of an uptake rate and the calculation of a BCF from biomagnification data (Crookes and Brooks 2011). Another approach similar to the Sijm approach has also been considered as potentially suitable. Barber et al. (2003) compared and reviewed models and methods for predicting bioconcentration in fish by comparing the predictability of these models using a set of experimental bioconcentration data. As a result an allometric equation, also independent of the log KOW, was derived using 517 data points. Using the elimination rate values for C11-PFCA measured by Martin et al. 2003b (please find details of the study in section 3.3.1.3)and taking into account growth i.e. increase of weight over time the calculated BCFs range between 5952-6326 for carcass and 5339-5675 for liver based on the allometric equation derived by Sijm et al. (1995). Based on the allometric equation from Barber et al. (2003), BCFs range between 5848-6071 for carcass and 5246-5675 for liver. The calculated BCFs from BMFs of the fish feeding study (Martin et al., 2003b) are larger than the measured BCFs (Martin et al., 2003a) but in a similar order of magnitude. The elimination rates were different in the two studies of Martin et al. (Martin et al., 2003a; Martin et al., 2003b), which may according to the authors be due to the different fish size. The BCF study (Martin et al., 2003a) gives measured uptake rates that are different than those estimated from the size of the fish. This indicates that the equations estimating uptake rates may not be correct for these compounds. Therefore, the use of the BCF calculations is rather speculative and is only used as additional evidence.

A recently published comparison of BCFs and biomagnification factors (BMFs) investigated 9 substances in a laboratory fish feeding study with carp (Inoue et al. 2012). Five substances showed BCFs larger than 5000 but only two of these substances were likely to biomagnify. Based on linear regression conducted with their data the authors suggest that a BMF of 0.31 indicates a high bioaccumulation potential. We used the linear regression to calculate BCFs based on BMFs of the study by Martin et al (Martin et al., 2003b) (please find details of the study in section 3.3.1.3). Again, the calculated BCFs from BMFs of the fish feeding study (Martin et al., 2003b) are larger than the measured BCF (Martin et al., 2003a) but in a similar order of magnitude and range between 4044 and 5132. However the results of this calculation may only be seen as equivocal evidence. Only nine data points were used to derive the linear regression and as the authors state themselves the results of the study are highly suggestive and more data would be necessary to support their findings.


23

The calculated kinetic BCFs are summarized in Table 8.

Table 7 measured Bioconcentration factors (BCF) of C11 –PFCA.

Species/foodweb BCF Reliability Reference

Rainbow trout (carcass) 2700 ± 400 2

Rainbow trout (blood) 11000 ± 1400

Rainbow trout (liver) 4900 ± 770

(Martin et al., 2003a)

Carp (whole) 2300 - 3700 1 (National Institute of Technology and Evaluation, 2007)

Table 8 Calculated BCFs based on elimination rate and BMF values for C11-PFCA. The BMFs were measured by Martin et al. 2003b (please find details of the study in section 3.3.1.3).

Species/foodweb BCF Reliability Reference

Juvenile rainbow trout (Carcass)

5952-6326 2

Juvenile rainbow trout (Liver)

5339-5675 2

(Sijm et al., 1995; Martin et al., 2003b)


5848-6071 2

Juvenile rainbow trout (Liver)

5246-5675 2

(Barber et al., 2003; Martin et al., 2003b)


4044-5132 3 (Inoue et al., 2012; Martin et al., 2003b)

Conclusion:

In general, all the BCFs reported indicate a high bioaccumulation potential of the C11-PFCA. The whole body BCF value is above the trigger value of 2000 indicating that C11-PFCA is bioaccumulative (B) but the trigger value below 5000 indicates that C11-PFCA is not a very bioaccumulative substance. Unlike lipophilic organic compounds C11-PFCA does not preferentially accumulate in adipose tissue but in protein rich tissues. Concentrations were highest in blood, kidney, liver and gall bladder and low in the gonads, adipose and muscle tissue. BCF values up to 12400 were achieved for blood, whereas the BCF-values up to 5670 were achieved for liver. A BCF of 3100 for carcass was reached as a maximum value. Taking into account the property of C11-PFCA to preferably bind to protein rich tissues such as liver, which may lead to a critical concentration in the liver and may cause hepatoxicity as has been shown for C10 and shorter chained PFCA, organ specific BCF value may have more importance from a toxicological point of view. However, conclusions on bioaccumulation should normally be based on whole body values and in this case, carcass is seen a good approximation for whole body by the authors.

BCF values were also calculated on the basis of biomagnification data by different methods. Two allometric approaches reviewed and considered appropriate for the calculation of BCFs from dietary study data yielded calculated BCF values larger than 5000 for either carcass or liver. Calculated values are in the same magnitude of order as experimental BCFs. However, some uncertainty remains as these calculations are rough estimates. BCF values calculated with the third method, based on a linear regression,are also in the same order of magnitude. However, the results from the third calculation should be regarded with caution as more data is needed to confirm this linear relationship.


24

Based on the BCF results there is sufficient evidence that C11-PFCA is a bioaccumulative(B) substance and likely a vB substance, but no final conclusion can be drawn on the vB properties.

3.3.1.2 Bioaccumulation factors (BAFs)

In field studies on bioaccumulation of chemicals bioaccumulation factors (BAF) are measured. The BAF is typically measured in the field as the ratio of the chemical concentrations in the organism and the water:

Water

Biota

c

cBAF =

where chemical concentration in the organism (CBiota) is usually expressed in units of gram of chemical per kilogram of organism. The weight of the organism can be expressed on a wet weight basis or appropriately normalized, if needed, (e.g.,lipid- or protein-normalized) (Conder et al., 2011). BCFs are measured under controlled laboratory conditions, whereas BAF is a field measurement and therefore different from BCF. Moreover, the chemical exposure is not only limited to water but all routes including diet.

Labadie et al. (2011) investigated the partitioning of various PFCAs in the Orge River, an urban tributary of the Seine River. Bioaccumulation and tissue distribution was studied in European chup, a common cyprinid in European freshwater and a benthopelagic fish. Five adult fish were collected in April 2010. The sex of each individual fish was analysed according to gonad morphology. Whole liver, gills and gonads were taken along with portions of muscle. Water and sediment samples were taken as triplicates at the same site. Large inter-individual variations, not sex-related, were observed. In agreement with the findings made by Martin et al. (2003a) tissue distribution shows that C11-PFCA is especially accumulated in blood and liver. The results of this study are summarized in table 9. All values are above 5000, although this trigger value relates to whole body BCFs. Thus this study suggests that C11-PFCA is very bioaccumulative.

Furthermore, BAFs were calculated from water and biota concentrations reported in the studies of Loi et al (2011) and Houde et al (2006) (see Table 9). The studies are described more in detail in sections 3.3.1.4 and 3.3.1.3, respectively. Variations in calculated BAFs originate from variations in measured concentrations in fish.

Table 9 Bioaccumulation factors (BAF) of C11 –PFCA in different aquatic foodwebs detailed in section 3.3.1.2

Species Log BAF BAF Reliability Reference

Chup(Leuciscus cephalus) plasma

6.0±0.1 1000000

Chup(Leuciscus cephalus) liver

5±0.3 100000

Chup(Leuciscus cephalus) gills

4.9±0.1 79433

Chup(Leuciscus cephalus) gonads

4.8±0.2 63096

Chup(Leuciscus cephalus) muscle

4.3±0.2 19953

2 Labadie et al. (2011)


25

Grey mullet (Mugil cephalus); n = 5

3.6 (mean) 3.7 (median)

4292 (mean) 5263 (median)

Mozambique tilapia (Oreochromis

mossambicus); n = 5



Small snakehead (Channa asiatica); n = 3


15417 (mean) 21842 (median)

Ladyfish (Elops saurus) n = 6



Flag-tailed flass perchlet (Ambassis

miops) n = 2

3.8 (mean) 6542 (mean)

2 Loi et al. (2011)

Striped mullet (Mugil cephalus)

3.4 (mean) 3.7 (max conc. fish to mean water)

2818 (mean) 5409 (max conc. fish to mean water)

2 Houde et al. (2006)

Pinfish (Lagodon rhomboids)



Red drum (Sciaenops ocellatus)



Atlantic croaker (Micropogonias undulates)



Spotfish (Leiostomus xanthurus)



Spotted seatrout (Cynoscion nebulosus)



In conclusion, the derived BAFs indicate that C11 –PFCA is very bioaccumulative.

3.3.1.3 Biomagnification factors (BMFs)

Besides bioconcentration also biomagnification describes the potential of a chemical to bioaccumulate. Biomagnification factors (BMFs) can be measured in the laboratory in a fashion similar to that used in the OECD and US-EPA bioconcentration test protocols. Organisms are exposed to a chemical preliminary via diet. The BMF test typically includes an uptake phase, where levels of chemicals are followed over time, ideally until the chemical concentration in the organism no longer changes with time (i.e., reaching the steady-state). If a steady-state cannot be reached in the experiment, the uptake phase is followed by a depuration phase where organisms are exposed to uncontaminated food. The rate of decline in chemical concentration over time measured in the depuration phase can then be used to derive the chemical uptake rate from which a hypothetical steady-state concentration can be estimated (Conder et al., 2011). The laboratory-derived dietary BMF is calculated using the ratio of the chemical concentrations in the test animals at steady-state and their diet:

diet

biota

C

CBMF =


26

where chemical concentration in the organism (Cbiota) and its diet (Cdiet) are appropriately normalized, if needed, (e.g., lipid- or protein-normalized) (Conder et al., 2011). BMF values based on field studies are based on the ratio of the concentration in the predator and the prey:

prey

predator

field C

CBMF =)(

There are several uncertainties concerning field based BMFs similar to field based trophical magnifacation factors which regard food webs. There are biological, ecological factors which can influence the outcome of a BMF. Additionally as there is no standard procedure so far how to conduct such field studies and therefore different study designs may too have an influence. The uncertainties of field studies have been addressed and discussed by Borga et al. (2012). As the authors actually refer to field based trophical magnifacation factors a summery of the discussion has been included in chapter 3.3.1.4 trophic magnifacation factors. Problems arise with increasing body size of predators because analysis is based on tissue or serum samples. Whole-body analysis is not feasible for ethical reasons, i.e. a whole whale would be needed, and due to the challenging logistics with respect to sampling and laboratory constraints. Therefore, some of the derived BMF-values are restricted to certain tissue samples rather than whole body samples. BMF values based on liver samples may be over estimative. From a toxicological perspective concentrations in individual organs, such as the liver, may be more relevant when the potential for direct organ-specific toxicity (i.e., liver toxicity) is being predicted. Whole body values may be estimated if the tissue mass fraction is known for the organism regarded. There may however be some uncertainties due to inter individual and geographical differences (Houde et al., 2006). Martin et al. (2003b) exposed juvenile rainbow trout (Oncorhynchus mykiss) for 34 days to PFCAs in the diet, followed by a 41 day depuration period. Though, the authors describe their results as BAF the results of this study should rather be assigned as BMFs according to the above mentioned definition as uptake only derived from the diet. During the uptake period, animals were daily fed with spiked food at a rate of 1.5 % food per body weight. Spiked food concentrations were 0.57 µg/g for C11-PFCA. Water samples collected before and after feeding revealed no traces of PFCAs in water. At 6 occasions during uptake period and during depuration period, fish were removed to determine the kinetics of uptake and depuration. The authors estimated the steady state to less than 34 days. Carcass and liver concentrations were determined by using liquid chromatography-tandem mass spectrometry, and kinetic rates were calculated to determine bioaccumulation parameters. Bioaccumulation (carcass) increased with increasing chain length but was not larger than one: 0.28 ± 0.04 for C11-PFCA; (see also Table 10). This indicates that a dietary exposure will not result in biomagnification in juvenile trout. The authors assume that the lack of observed biomagnification was likely due to the small size of fish used in the study, resulting in more rapid chemical elimination to water, relative to body size and and that their natural feeding rate is too low. This more rapid chemical elimination would reduce the BMF stronger than what would be observed for larger species or size classes (Martin et al., 2003b).

Furthermore BMFs were estimated from field studies. Studies are described below and results are shown in Table 10.

Transfer of PFCAs was elucidated in Lake Ontario including one 4-membered pelagic food chain (Martin et al., 2004). Whole body samples were collected. Two macroinvertebrates (Diporeia and Mysis) were considered as primary prey whereas rainbow trout inhabited the top predator´s position. Lake trout samples were taken at various locations and years (1980-2001) in Lake Ontario. Seven samples were selected every three years (i.e. 7 individual fish samples per year). Forage fish species, including sculpin, smelt, and alewife, were collected on October 9th 2002 at an offshore site near Niagara-on-the-Lake, Lake Ontario. .Due to the inherent uncertainties correlated with constitution of diet 4 individual combinations of rainbow trout and its prey were regarded. As this study was conducted with fish uptake of PFCAs may not have occurred exclusively over diet but also over the gills. Thus, the factors may be more accurately addressed as BAF. A striking finding of this study was the unexpectedly high content


27

of PFCAs in the benthic invertebrate Diporeia occupying the lowest trophic level. The mechanism leading to this exceptional accumulation still needs to be unravelled. The author’s hypothesis is that sediments are a major source for PFCAs. Results are given in Table 10. Tomy et al. also investigated liver samples of the beluga whale, ringed seal, fish pelagic and whole body samples of amphipod and arctic copepod of the Western Canadian Arctic. The animals selected were from the sample archived repository at Fisheries and Oceans, Canada. Blubber and liver of beluga (n = 10, all males,) from Hendrickson Island and ringed seal (n = 10, all males) from Holman Island were collected in 2007 and 2004, respectively. Fish species collected in 2004 and 2005 included the marine pelagic Arctic cod (n = 10) from the Amundsen Gulf, the marine coastal Pacific herring (n = 10) from the Mackenzie Shelf and the anadromous Arctic Cisco (n = 9) from the Mackenzie estuary. The marine pelagic amphipod Themisto

libellula (pooled samples, n = 2) and the marine Arctic copepod Calanus hyperboreus (pooled samples, n = 5) were collected in 2004 from the eastern Beaufort Sea and Amundsen Gulf region. As the authors state themselves differences in sampling years may influence the interpretation of the food web transfer. On the other hand the Arctic as a remote area may be less prone to temporal changes and the existence of point sources there is unlikely. The derived BMF-values (see Table 7 and Figure 1) are restricted to the liver and the resulting BMF may be over estimative. From a toxicological perspective concentrations in individual organs, such as the liver, may be more relevant when the potential for direct organ-specific toxicity (i.e., liver toxicity) is being predicted (Tomy et al., 2009). Houde et al. examined PFCA serum concentrations in bottlenose dolphins at two different habitats, Charleston Harbor and its tributaries (i.e., the Cooper, Ashley, and Wando rivers) and the Stono River estuary, South Carolina, and in Sarasota Bay, Florida.. Marine water (n=18), surface sediment (n=17), Atlantic croaker(n=3), pinfish(n=4), red drum (n=8), spotfish (n=10) , spotted seatrout (n=11), striped mullet (n=8), and bottlenose dolphin samples (n=24) were collected around the Charleston Harbor area. Marine water(n=10), surface sediment (n=8), zooplankton 8n03), sheephead (n= 3), pigfish (n= 10), pinfish (n=10), striped mullet (n=9), spotted seatrout (n=8), and bottlenose dolphin samples 8n=12) were collected at Sarasota Bay. Dolphin plasma, skin, and teeth were collected from both locations. Additionally, dolphin tissue samples (i.e., liver, kidney, muscle, lungs, heart, thyroid, and thymus) were collected of recently deceased bottlenose dolphins from Sarasota Bay (2002, n = 1, male, 233.5 kg) and Charleston (2003, n = 1, female, 708.4 kg). Additional liver (n = 6) and kidney (n = 6) samples collected from stranded bottlenose dolphins were available at Sarasota Bay. The authors claim that utilization of serum or liver concentrations of dolphins will overestimate the BMF by a factor of 10 - 30. Whole body concentrations were estimated on the basis of tissue distribution. In the course of the study PFCA serum concentrations in bottlenose dolphins were examined at two different habitats. Samples were collected between 2002 and 2004. Unfortunately, concentrations in other representative fish species originated from different years, thus, entailing additional uncertainty when assessing BMF through the food chain. Wastewater treatment plant discharges in the Charleston area may have resulted in non–steady state concentrations in the food web. On the other hand it may be assumed that media and biota were continuously exposed to PFCA in this area throughout the years. The results are summarized in Table 10 (Houde et al., 2006). Butt et al. conducted a study in the Canadian Arctic. Ringed seal liver samples (n=10 per site) were provided by local hunters from 11 different locations in the Canadian Arctic. The age of the animals was determined via tooth aging and for a few samples the age was estimated using length-age correlations. Stable isotope analysis was done with 15N to 14N and 13C to 12C. Based on liver samples from polar bears obtained from another study and ringed seal data measured in this study BMFs were calculated (see Table 8 and Figure 1). Mean concentrations in the liver ranged between 2.1-13.3 µg/g wet weight. The polar bear sample sites were associated with ringed seal populations. However, the sample collection year for ringed seal populations varied from 2002 to 2005, and it is possible that interpretation of spatial trends may be confounded by temporal variations of PFCA concentration within seal populations (Butt et al., 2008). Various predator prey relationships in the Westerschelde (Netherlands) were investigated by van Heuvel-Greve and co-workers. Samples of habour seal plasma (n=3) and whole body


28

samples of herring, seabass and flounder (n=4 per species) as well as zooplankton were collected in 2007 and 2008. The trophic level was estimated based on stable isotope (15N) analysis. BMFs were considerable for harbour seal as well as for the sediment dwelling flounder (see Table 10 and Figure 1) (Environment Canada, 2012; van den Heuvel-Greve et al., 2009).

Table 10: Biomagnification factors (BMF) for C11-PFCA; if not indicated otherwise BMFs refer to whole body

Location Species/foodweb BMF Reference Relaibility

Canadian Arctic Polar bear (liver)/ ringed seal (liver) 7.1-21 Butt et al. 2008 2

US, South Carolina, Charleston

Seatrout /pinfish 0.9

US, South Carolina

Dolphin (whole, estimated)/striped mullet

1.9

US, South Carolina

Dolphin (whole, estimated)/pinfish 2.4

US, South Carolina

Dolphin(whole, estimated)/red drum 3.2

US, South Carolina

Dolphin (whole, estimated) /atlantic croaker

2.1

US, South Carolina

Dolphin (whole, estimated) /spotfish 3.9

US, South Carolina

Dolphin (whole, estimated)/seatrout 2.5

Houde et al. 2006

2

Western Canadian Arctic

Ringed seal (liver)/ arctic cod (liver) 6.6


Beluga whale (liver)/ arctic cod (liver) 229


Beluga whale (liver)/ Pacific herring (liver)

353


Beluga whale (liver)/ arctic cisco (liver)

181

Tomy et al. 2009 2


29


Arctic cod (liver)/ marine arctic copepod (whole body)

0.3


Arctic cod (liver)/ marine pelagic amphipod (whole body)

0.3

Lake Ontario Lake trout/alewife 6.4

Lake Ontario Lake trout/smelt 1.2

Lake Ontario Lake trout/sculpin 0.21

Lake Ontario Lake trout/prey (weighted) 3.4

Martin et al. 2004

2

Laboratory juvenile rainbow trout(Carcass) 0.28±0.04 Martin et al. 2003b

2

Herring/ Zooplankton 1.9

Herring/ sea bass 3.2

harbour seal/herring 53

harbour seal/sea bass 17

peppery furrow shell/flounder 10

lugworm/flounder 25

Westerschelde, Netherlands

flounder/habour seal 9.0

(Environment Canada , 2012; van den Heuvel-Greve et al., 2009)

2


30

Comparing the different BMF values for different PFCAs within a single study revealed a trend between chain-length and BMF-values. Highlighted in figure 1 are the BMF values from ringed seal/polar bear for C9-C14 showing that C11-PFCA has a higher biomagnification potential than the longer chained homologues. This indicates that C11-PFCA has a higher potential to

biomagnify in predators than the longer chained PFCA. (See figure 1)

0

20

40

60

80

100

120

C9-PFCA C10-PFCA C11-PFCA C12-PFCA C13-PFCA C14-PFCA

BM

Fs

Figure 1: Biomagnification factors (BMFs) for C9-14-PFCA (Butt et al. 2008).

Conclusion:

The biomagnification potential of C11-PFCAs was investigated in several field studies and one laboratory study. As discussed, several uncertainties can influence the results of field data. Nevertheless, the overall weight of evidence is conclusive. Various field studies investigating the biomagnification potential between different predator/prey-relationships showed BMFs well above one indicating biomagnification. Biomagnification was greater in homeotherms than in poikilotherms. For dolphin, polar bear beluga whale and ringed seal, BMFs of several magnitudes greater than one have been reported. Comparing the different BMF values for different PFCAs within a single study revealed a trend between chainlength and BMF-values, showing that C11-PFCA has a higher biomagnification potential than the longer chained homologues. On the other hand the longer chained PFCA show higher bioaccumulation potential referred to the “classical” BCF and can be assigned as very bioaccumulative (see Data Matrix in Annex I and figure 1) based on the BCF data alone.

3.3.1.4 Trophic magnification factors (TMFs)

The trophic magnification factor (TMF) is a measure to evaluate biomagnification occurring in food webs. In the Guidance Document on Information Requirements, Chapter R.7.10.1.1, TMF is defined as the concentration increase in organisms with an increase of one trophic level. Again a TMF greater than one indicates accumulation within the food chain.


31

There are several uncertainties concerning TMFs. These have been addressed and summarized by Borga et al. (2012). There are biological factors such as the differences between poikilotherms and homeotherms, sex, different energy requirements, different abilities to metabolize chemicals and slow or fast growing organisms. Steady state between a consumer and its diet is assumed. However, as opportunistic feeders wild animals vary their diet over seasons or with life stage and point sources may influence observed TMFs. Additionally, apart from the diet there always is the possibility of a direct uptake of the substance under scrutiny and the relative importance of food versus e.g. water exposure can influence the magnitude of the TMF. The position in the food web is quantified using relative abundances of naturally occurring stable isotopes of N (15N/14N, referred to as δ15N). However the relative abundance of these isotopes und thus the determination of the trophical level and TMF is influenced by the physiology of the organism and its life trait history. Rapid growth with a higher protein demand for new tissue leads to lower enrichment factors than those with slower growth rates. Insufficient food supply and fasting and starvation leads to catabolism of body proteins and an increase of 15N in organisms relative to those organisms with adequate food supply. There is no standard procedure for the conductance of TMF field studies. Hence, the conductance and sampling may vary between different studies. Disproportionate sampling of the food web or unbalanced replication of samples may significantly influence the TMF. Additionally, as already discussed in the BMF chapter sample collection is often restricted to tissue or serum samples with increasing body size of predators due to ethical reasons and due to the challenging logistics with respect to sampling and laboratory constraints.

Martin et al. examined PFCA contents in the food web from Lake Ontario in Canada (Martin et al., 2004). Adult lake trouts (top predator) were collected at various years and locations in Lake Ontario. Samples of prey fish (sculpins, smelts and alewifes) and macroinvertebrates (Mysis sp., Diporeia sp.) were collected at one location in October 2002. A more detailed study description is given in chapter 3.3.1.3. Lake trout samples analyzed represented individual whole fish homogenates. The other species were processed as composites of whole individuals. TMFs are shown in Figure 2 and Table11.

Houde et al. investigated the food web of bottlenose dolphins. The authors sampled different biota, i.e. croaker, pinfish, spotfish, spotted seatrout, striped mullet and samples from bottlenose dolphins, as well as water and surface sediment. Sample collection was conducted between 2002 and 2004. Based on stable isotope (15N) analysis the trophic level of each biota sample was determined. PFCAs were analysed in plasma and liver of dolphins and afterwards a whole body burden was calculated. For prey whole body homogenates were analysed for PFCA (Houde et al., 2006). A more detailed study description is given in chapter 3.3.1.3. For results see Table 11.

Kelly et al. measured PFCAs in the Canadian Arctic marine food web. The authors used concentrations in sediment and in different organisms (lichens, macroalgae, bivalves, fish) and tissues and organs (stomach contents, liver, muscle, blubber and/or milk) of common eider ducks, and beluga whales and ringed seals to calculate TMFs. Sample collection was conducted between 1999 and 2003 along the eastern Hudson Bay coastline in close proximity to the Inuit village Umiujaq. PFCAs were measured in different tissues/fluids of the beluga whale including blood, muscle liver, milk and also in foetuses. The authors could show that PFCAs especially accumulate in protein rich compartments such as blood and liver and that the TMF of PFCAs correlates with the partitioning behaviour between protein and water and protein and air. Comparisons of different food webs show that the TMF is below one in the case of piscivorous food webs if air breathing organisms are excluded but becomes larger than one if air breathing organisms are taken into account (see Table 11 and Figure 2) (Kelly et al., 2009).

Loi et al. investigated a subtropical food web in a nature reserve including phytoplankton, zooplankton, gastropod, worm and shrimp, and liver samples of fish, and water bird. Samples were collected between 2008 and 2010. Surface water (n =12), sediment (n = 6), phytoplankton in general [pooled samples (p) = 3], zooplankton [mainly amphipods and


32

copepods (p = 2)], gastropods (p = 3), worms (p = 10 of 3 families: Capitellidae, p = 5; Nephtyida, p = 3; Sabellidae, p=2), shrimps [p = 4 of 2 species: black tiger prawn (Penaeus monodon), p = 2; sand prawn (Metapenaeusensis), p = 2], fish [n=21 of five species: greymullet (Mugil cephalus), n = 5; ladyfish (Elops saurus), n = 6; Mozambique tilapia (Oreochromis mossambicus), n = 5; small snakehead (Channa asiatica), n =3; flag-tailed glass perchlet (Ambassis miops), p = 2 with each pool consisting of 27 individuals] were collected from a tidal shrimp pond. Surface water and sediment samples were collected concurrently with the biota samples. BSAF were calculated based on the assumption that sediment was the major exposure pathway for worms. The study investigated PFCAs with different chain length. C11-PFCA was detected in sediment only (Loi et al., 2011).

Table 11: Trophic Magnification Factors (TMF) of C11-PFCA; if not indicated otherwise TMFs refer to whole body.

Location Species/foodweb TMF Reference Reliability

US, South Carolina Dolphin plasma, croaker, pinfish, spotfish, spotted seatrout

3.0 ± 3.9 2

US, South Carolina, Whole dolphin burden, croaker, pinfish, spotfish, spotted seatrout

2.3 ± 2.5

Houde et al. 2006


western arctic food web; beluga whale, ringed seal, fish, amphipod, arctic copepod

5.9 - 31.2 Tomy et al. 2009

2

Hudson Bay (north-eastern Canada

Sediment/ macroalgae/ bivalves/ fish/ seaduck/ beluga whale

6.25 - 10.2 3.63 - 6.32 (protein corrected)

2

Hudson Bay (north-eastern Canada

Sediment/ macroalgae/ bivalves/ fish

0.75 - 1.58 (protein corrected)

Kelly et al. 2009

Lake Ontario Diporeia/slimy sculpin No significant association with trophic level

2

Lake Ontario Mysis/alewife/rainbow smelt/lake trout

4.71

Martin et al. 2004

Mai Po Marshes Nature Reserve in Hong Kong

Tidal shrimp pond brackish food web

1.74 Loi et al., 2011

2


33

0

1

2

3

4

5

6

7

8

9

C9-PFCA C10-PFCA C11-PFCA C12-PFCA C13-PFCA C14-PFCA

TM

Fs

Kelly et al., 2009

M artin et al. 2004

Figure 2: Trophic Magnification Factors (TMF) of C9-14-PFCAs.

Equally to the findings for BMF, comparing the different TMF values for different PFCAs within a single study revealed a trend between chain-length and TMF-values, showing that C11-PFCA has a higher biomagnification potential than the longer chained homologues. The reduced biomagnification of the longer chained PFCA may be due to reduced bioavailability of the longer chained PFCAs caused by association with dissolved an colloidal organic matter (Kelly et al., 2009; Martin et al., 2003a).

Conclusion:

A number of field studies are available which analyzed the trophic magnification potential of C11-PFCA. Several uncertainties and different test designs of the field studies may influence the outcome and may lead to contradictory results. Nevertheless, considering the potential to biomagnify in a trophic food web the overall outcome is conclusive. For food chains of dolphin, beluga whale, and ringed seal, TMFs greater than one ranging between 2.3 and 31.2 have been reported, indicating high trophic biomagnification. Trophic magnification was greater if the food chain contained homeotherms. TMFs were smaller in the case of piscovorous food webs and if air breathing organisms are excluded but became larger if air breathing organisms were taken into account (Kelly et al., 2009). Thus, C11-PFCA biomagnifies in some food chains analyzed within the food web. Different TMF values for different PFCAs within a single study revealed a trend between chain length and TMF-values, showing that C11-PFCA has a higher biomagnification potential than the longer chained homologues. The decline of the potential to magnify in the food web from C11-PFCA to C14-PFCA indicates that trophic magnification is more pronounced for C11-PFCA than for the longer chained PFCAs. On the other hand the longer chained PFCA show higher bioaccumulation potential referred to the “classical” BCF and can be assigned as very bioaccumulative (see Data Matrix in Annex I and figures 1 and 2).

3.3.2 Terrestrial bioaccumulation

Müller et al. conducted a terrestrial food web study consisting of lichen and plants, caribou, and wolves from two remote northern areas in Canada. Liver, muscle, and kidney samples

(n=7 Porcupine herd food web and n=10 for the Bathurst food web) from two caribou herds were collected; from the Porcupine herd in northern Yukon Territory and the Bathurst herd in the Northwest Territories (NWT)/western Nunavut . Wolf (n=6 Porcupine herd food web and


34

n=10 for the Bathurst food web) lichen, and plant samples were collected in the same region as the caribou. Plant samples included cottongrass, aquatic sedge, willow, moss, and mushrooms. Liver and muscle samples were collected from the sampled wolves. Lichen, moss and mushrooms were collected as a whole grass and willow without roots. Plant samples are from the same season (Summer 2008 in Porcupine or Summer 2009 in Bathurst) whereas wolf and caribou samples are from different years (2007 and 2010 in Porcupine and 2008 and 2007 in Bathurst). This food web is considered as relatively well documented example (Kelly and Gobas, 2003). The study illustrates a considerable carry over between plants and caribou. Caribou are a major human food source in numerous arctic communities. This food-chain may also be considered comparable to the pasture-cow food-chain in temperate regions. The results of the study, BMFs as well as TMFs are shown in table12 and table 13. Tissue concentrations and whole body concentrations were used for calculations. Tissue based BMFs differ considerably. Therefore it is concluded that BMFs based on whole body concentrations are more appropriate. Whole body concentrations were estimated (Müller et al., 2011).

Table 12: TMF values for C11-C13-PFCA for a terrestrial food chain (Müller et al., 2011).

PFCA BMF

Caribou (whole)/lichen

BMF

Caribou (whole)/vegetation

BMF

Wolf (whole)/caribou (whole)

Porcupine Bathurst Porcupine Bathurst Porcupine Bathurst

C11 3.8± 0.9 8.2 ± 1.1 9.8± 3.2 14.5 ± 1.8 2.0± 0.3 2.8 ± 0.5

C12 2.9± 1.1 11± 3 4.5± 1.7 8± 3 1.2± 0.1 1.4± 0.4

C13 2.4± 1.0 6.9± 2.2 7.1±4.7 9.0±2.5 0.8± 0.2 3.2± 1.5

0

2

4

6

8

10

12

14

16

C9-PFCA C10-PFCA C11-PFCA C12-PFCA C13-PFCA

BM

Fs

(Porcupine)Caribou

(whole)/lichen

(Porcupine) Caribou

(whole)/vegetation

(Bathurst) Caribou

(whole)/vegetation

Figure 3: BMFs for C11-13-PFCAs in a remote terrestrial food chain from two different locations (whole body, Müller et al., 2011). The study is reliable (reliability 2).

Table 13: TMF values for C11-C13-PFCA for a terrestrial food chain (Müller et al., 2011).

PFCA TMF

Wolf (liver) /caribou

TMF

Wolf (whole)/caribou

TMF

Wolf (whole)/caribou


35

(liver)/lichen (whole)/lichen (whole)/vegetation

Porcupine Bathurst Porcupine Bathurst Porcupine Bathurst

C11 6.6± 0.5 6.1 ± 0.3 2.5± 0.2 2.8 ± 0.1 2.2± 0.2 2.9± 0.3

C12 4.1± 0.1 5.2± 0.3 1.4± 0.1 2.2 ± 0.1 1.3± 0 2.0± 0.1

C13 3.7± 0.1 4.2 ± 0.3 1.4± 0.1 2.0 ± 0.1 1.4± 0.0 1.8± 0.1

0

1

2

3

4

5

6

7

8

C9-PFCA C10-PFCA C11-PFCA C12-PFCA C13-PFCA

TM

Fs

Wolf (liver) /caribou (liver)/lichen

(Porcupine)

Wolf (liver) /caribou (liver)/lichen1

(Bathurst)

Figure 4: TMFs for C11-13-PFCAs in a remote terrestrial food chain from two different locations (whole-body, Müller et al., 2011).

For comparison and to highlight the observed trend all BMF and TMF values of the PFCA with different chain length are given in table 11 and table 12. Data also shown in figure 3 and figure 4 are bold in the tables. Observations made in the terrestrial food web confirm the observations made in aquatic food webs. C11-PFCA has a higher potential to be transferred through the food chain than its longer chained homologues..

Conclusion:

The terrestrial BMF and TMF of C11-PFCAs are greater than one for the remote Arctic food chain lichen – caribou – wolf, ranging between 2 and 14.5 for BMFs based on estimated whole body burdens and ranging between 2.2 and 2.9 for TMFs based on estimated whole body burdens. TMFs and BMFs show clearly higher values for C11-PFCA compared to C12,13-PFCAs (see Data Matrix in Annex I and figures 1-4).This confirms the findings from aquatic food webs. In Addition the following should be taken into account. Caribou are a major human food source in numerous arctic communities and this food-chain may also be considered comparable to the pasture-cow food-chain in temperate regions. Hence, C11-PFCA could become a concern for humans in different climate zones.

3.3.3 Monitoring

There are several monitoring studies. However these will not be all addressed here and focus is put unto one study conducted by Freberg et al. (2010) which supports the findings for BMFs and TMFs. PFCAs have been determined in the serum of professional ski waxers (Freberg et al. 2010). Ski waxers are known to be highly exposed to PFCAs (Nillson, 2010). Thirteen ski waxers participated in the study. They were 28 – 53 years old and have been working as a ski waxer for 2 – 13 years. Blood sample were collected at the end of one season and the beginning and end of the next season. Instrumental analysis was performed with liquid


36

chromatography coupled to a triple quadropole mass spectrometer. C4-14-PFCAs have been analysed, whereby C5-PFCA was not detected. Limit of quantification was 0.05 ng/ml for all PFCAs (except of C4-PFCA 0.1 ng/ml). The reductions in the concentrations measured at the start of the second season compared to the end of the first season were all of statistical significance (p < 0.05), except for PFNA (p = 0.29). The concentrations of the C8-11-PFCAs were reduced by five up to 20% on average during the eight month period without occupational exposure to ski waxes. The reduction was substantially larger for the other determined PFCAs, i.e. > 30 % for C12,13-PFCAs and > 60 % for C14-PFCA.

The results indicate that C11-PFCA is eliminated from human serum at a slower rate compared to C12-14-PFCAs. Therefore the results further support what has been seen from comparison of TMFs and BMFS, which is that C11-PFCAs shows a higher bioaccumulation potential compared to C12-14-PFCAs in certain studies.

3.3.4 Summary and discussion of bioaccumulation

The whole body BCF for C11-PFCA from the MITI study conducted with carp is in the range of 2300 – 3700 (National Institute of Technology and Evaluation, 2007). The study of Martin et al., shows that carcass BCFs closely approximate the whole-body BCF. This means the whole-body BCF would be approximately 2700 (mean value). C11-PFCA would therefore be bioaccumulative but not very bioaccumulative if referred to whole body according to Annex XIII of REACH.

Taking into account that PFCAs show hepatotoxic effects, focus could also be put on the higher accumulation potential in liver as bioaccumulation may be used as indicators for toxicity to organisms. Based on the findings there is a potential to reach critical levels that may elicit toxic effects over long-term exposures due to the high bioaccumulation potential to the liver.

In conclusion, based only on the BCFs C11-PFCA is a bioaccumulative substance (B) but a borderline substance for vB, where it cannot be derived for sure whether the vB criterion of REACH is fulfilled or not.

However, calculated BCF values on the basis of biomagnification data are larger than 5000 for either carcass or liver and support that C11-PFCA is very bioaccumulative. Additionally, field based studies provide BAF values for C11-PFCA well above the trigger-value of 5000 for various species and body compartments.

Additionally, if taking into consideration BMFs and TMFs it can not only be additionally concluded that C11-PFCA- biomagnifies as indicated by BMFs and TMFs but also that trophic magnification is more pronounced for C11-PFCA than for the longer chained PFCAs. Comparing the different BMF and TMF values for different PFCAs within a single study revealed a trend between chain length and BMF and TMF -values, showing that C11-PFCA has a higher biomagnification potential (with very high BMF values ranging between 0.2 and 353) than the longer chained homologues (see Data Matrix in Annex I and Figures 1-4). The potential to magnify in the food web declined from C11-PFCA to C14-PFCA indicating that trophic magnification is more pronounced for C11-PFCA than for the longer chained PFCAs. Hence, C11-PFCA has a lower bioaccumulation potential than its longer chained homologues but a higher potential to transfer through the food web.

The reduced biomagnification of the longer chained PFCA may be due to reduced bioavailability of the longer chained PFCAs caused by association with dissolved an colloidal organic matter (Kelly et al., 2009; Martin et al., 2003a).

A human biomonitoring study indicates that C11-PFCA is eliminated from human serum at a slower rate compared to its longer chained homologues C12-14-PFCAs. These results further support that C11-PFCAs shows a higher bioaccumulation potential compared to C12-14-PFCAs in certain studies.


37

There is not one key study which can prove alone that C11-PFCA fulfils the vB criteria. However, taking together all information as outlined above it can be concluded that C11-PFCA fulfils both the B and the vB-criteria of REACH..

This conclusion is also supported by the Canadian assessment of PFCAs (Environment Canada, 2012).

4 Human health hazard assessment


5 Environmental hazard assessment


6 Conclusions on the SVHC Properties

6.1 PBT, vPvB assessment

6.1.1 Assessment of PBT/vPvB properties – comparison with the criteria of Annex XIII

A weight of evidence determination according to the provisions of Annex XIII of REACH is used to identify the substance as vPvB. All available information (such as results of standard tests, monitoring and modelling, information from the application of the category and analog approach (grouping, read-across) and (Q)SAR results) was considered together in a weight of evidence approach. The individual results have been considered in the assessment with differing weights depending on their nature, adequacy and relevance. The available results are assembled together in a single weight of evidence determination.

Persistence

For C11-PFCA no experimental data on degradation are available. Therefore, data from chemically similar compounds should be considered in a read-across approach (please see Annex 1 for further details). The degradation potential of substances differing only in the number of carbons in the perfluorinated carbon chain has been analyzed in some studies which indicate that long chain PFCAs are resistant to degradation in the environment.

PFCAs are synthetic compounds which contain a structural feature: a perfluorinated carbon chain combined with a carboxylic group. The perfluorinated carbon chain is a synthetic feature, there are no natural sources known. The chemical structure of these compounds differs only in the number of perfluorinated carbons in the carbon chain.

The stability of organic fluorine compounds has been described in detail by Siegemund et al., 2000: When all valences of a carbon chain are satisfied by fluorine, the zig-zag-shaped carbon skeleton is twisted out of its plane in the form of a helix. This situation allows the electronegative fluorine substituents to envelope the carbon skeleton completely and shield it from chemical attack. Several other properties of the carbon-fluorine bond contribute to the fact that highly fluorinated alkanes are the most stable organic compounds. These include polarizability and high bond energies, which increase with increasing substitution by fluorine. The influence of fluorine is greatest in highly fluorinated and perfluorinated compounds.


38

Properties that are exploited commercially include high thermal and chemical stability (Siegemund et al., 2000).

Comparing the physico-chemical properties of C8-14-PFCAs it becomes obvious that with increasing chain length water solubility decreases and the sorption potential increases. This trend is based on the increasing number of CF2-groups in the molecular structure. The molecular reason for the persistence of highly fluorinated chemicals is the shielding effect of the substituted fluorine atoms described by Siegemund et al., 2000. Thus, using the described read across approach, we conclude that C13-PFCA is a very persistent synthetic compound which is resistant to abiotic and biotic degradation and fulfils both, the P and the vP criteria of Annex XIII.

Abiotic degradation

The data on the three CF2-groups C8-PFCA indicate that abiotic degradation in the atmosphere is expected to be slow (atmospheric lifetime = 130 days). The hydrolytic half-life of C8-PFCA at 25°C is greater than 92 years, with the most likely value of 235 years under relevant environmental conditions (3M Co., 2001a). No photodegradation of C8-PFCA has been observed in studies conducted under relevant environmental conditions. The estimated DT50 for indirect photolysis is 349 days.

Biotic degradation

Standard screening tests are available for C8,9,12,14-PFCAs. No biodegradation at all has been detected for C9,12,14-PFCAs within 28 days. For C8-PFCA test results differ from “no biodegratation” to 13% biodegradation of the ammonium salt. Thus, it can be concluded that C8,9,12,14-PFCAs are not readily biodegradable.

For C8-PFCA a non-standard aerobic biodegradation simulation test, one non-standard anaerobic biodegradation simulation test and field monitoring data from contaminated sites provide evidence that no biodegradation in water, soil and sediment occurs.

Conclusion

Henicosafluoroundecanoic acid (C11-PFCA) has no degradation studies available.

Read across approach within C8-C14-PFCAs can be applied for the persistence assessment of these substances. C8-14-PFCAs contain a highly similar chemical structure, a perfluorinated carbon chain and a carboxylic acid group. The compounds differ only in the number of CF2-groups. As a result of comparing the experimental and estimated physico-chemical data of C8-PFCA (the analogue substance) with experimental and estimated data on C11-14-PFCAs it can be assumed that with increasing chain length water solubility decreases and the sorption potential increases (See Table 15 of the support document). It can be with a sufficient reliability stated that the behaviour of these chemicals follow a regular pattern.

Due to both structural similarity and a regular pattern of physico-chemical properties, C8-14-PFCAs may be considered as a group or a category of substances for the purpose of the PBT/vPvB assessment and the read-across approach can be applied within this group.

In general, the persistence of C11-C14-PFCAs can be explained by the shielding effect of the fluorine atoms, blocking, e.g., nucleophilic attacks to the carbon chain. High electronegativity, low polarizability and high bond energies make highly fluorinated alkanes to the most stable organic compounds. It is not expected that the carboxylic group in PFCAs alters this persistence of these chemicals. This fact is confirmed by a hydrolysis study which obtained a DT50 of >92 years for C8-PFCA in water. Screening studies of C8,9,12,14-PFCAs showed no biodegradation within 28 days. Non-standard abiotic degradation tests with C8-PFCA could not detect any degradation products under environmentally relevant conditions. Furthermore, screening biodegradation studies on C8,9,12,14-PFCAs and one non-standard anaerobic biodegradation simulation test with C8-PFCA provide evidence of high persistence. Additionally, elements of non-standard higher tier aerobic biodegradation studies on C8-PFCA provide further support that no biodegradation in water, soil and sediment occurs.


39

Therefore, based on the information summarized above, it is concluded that C11-PFCA is not degraded in the environment and thus fulfils the P- and vP- criteria in accordance with the criteria and provisions set out in Annex XIII of REACH.

6.1.2 Bioaccumulation

Bioaccumulative substances are defined in Annex XIII with a BCF >2000. If substances have a BCF >5000 they fulfil the criteria of being very bioaccumulative.

Results from bioconcentration and bioaccumulation studies in aquatic species:

The whole body BCF for C11-PFCA from the MITI flow-through study conducted with carp is in the range of 2300 – 3700 (National Institute of Technology and Evaluation, 2007). The study of Martin et al., propose that carcass BCFs closely approximate the whole-body BCF. This means the whole-body BCF would be approximately 2700 (mean value) C11-PFCA would therefore be bioaccumulative but not very bioaccumulative if referred to whole body according to Annex XIII of REACH.

However, if taking into account that PFCAs show hepatotoxic effects, focus could also be given to the higher accumulation potential in liver as bioaccumulation may be used as indicators for toxicity to organisms. Based on the findings there is a potential to reach critical levels that may elicit toxic effects over long-term exposures due to the high bioaccumulation potential to the liver.

BCF values of 5848-6326 (carcass) and 5246-5675 (liver) derived with two calculation methods from an experimental fish feeding study suggest that C11-PFCA is very bioaccumulative.

Other information on the bioaccumulation potential and ability of the substance to biomagnify

in the food chain:

Available field based bioaccumulation data show BAF values for C11-PFCA well above the trigger-value of 5000 for various species from different studies .

Various available field-BMFs and even TMFs for C11-PFCA provide evidence on that biomagnification of the substance takes place in nature between different trophic levels of food chains and from the bottom to the top of food chains.

Comparing the different field-BMF and TMF values for C9-14-PFCAs within single studies revealed a trend between chain length and BMF and TMF values, showing that C11-PFCA has a higher biomagnification potential than the longer chained homologues (see Data Matrix in Table 16 of the support document and Figures 1-4). The potential to magnify in the food web declined from C11-PFCA to C14-PFCA indicating that trophic magnification is more pronounced for C11-PFCA than for the longer chained PFCAs. Hence, although C11-PFCA seems to have in laboratory studies a lower bioaccumulation potential than its longer chained homologues, based on field data it has a higher potential to transfer through the food web as demonstrated by high BMF, TMF and BAF values. This observation especially points to the need of expert judgement based on weight-of-evidence.

A human biomonitoring study indicates that C11-PFCA is eliminated from human serum at a slower rate compared to its longer chained homologues C12-14-PFCAs. These results further support that C11-PFCA shows a higher bioaccumulation potential compared to C12-14-PFCAs in certain studies.

Conclusion on the bioaccumulation potential:

Based on only the BCFs derived from flow-through tests, C11-PFCA clearly fulfils the B criterion but some uncertainty remains whether it fulfils the vB criterion. BCFs derived from BMFs of a fish feeding study are above 5000. Although these laboratory data indicate that the vB criterion may be fulfilled, further confirmation is sought from field data.


40

Bioaccumulation factors derived in field studies are clearly above the trigger of 5000 which can be considered analogous to a BCF > 5000. Furthermore, the various available field-BMFs and TMFs for this substance provide evidence that biomagnification at high level takes place in nature.. It is important to note that among C11-14-PFCAs, C11-PFCA based on field-BMFs and TMFs has the highest bioaccumulation potential indicating that trophic magnification is more pronounced for C11-PFCA than for the longer chained PFCAs. C12-14-PFCAs fulfill the vB criterion. Although for a large part of available information direct comparison with the vB criterion cannot be made, this information and especially the field data provide clear evidence on that the substance behaves in a way corresponding to bioaccumulation behaviour of substances which by direct comparison of the data to the vB criterion meet the criterion. Consequently, it is concluded that C11-PFCA fulfils both the B and the vB-criteria of REACH.

6.1.3 Toxicity


6.1.4 Summary and overall conclusions on the PBT, vPvB properties

In conclusion, C11-PFCA is identified as a vPvB-substance according to Art. 57 (e) of REACH following the application of a weight of evidence determination using expert judgement by comparing all relevant and available information listed in Section 3 of Annex XIII of REACH with the criteria set out in Section 1 of the same Annex.

6.2 CMR assessment


6.3 Substances of equivalent level of concern assessment.



41

Annex I - Read-across approach

In general, the read-across approach can be applied if substances whose physicochemical and/or toxicological and/or ecotoxicological properties are likely to be similar or follow a regular pattern as a result of structural similarity. Those substances may be considered as a group or a category of substances. According to ECHA`s practical guide 6 “How to report read-across and categories” similarities may be due to a common functional group, common precursor or breakdown products, constant pattern in changing potency or common constituents or chemical class.

Structural similarities of C8-14 PFCAs and some salts

In Table 14 the chemicals structures of the C8-14-PFCAs are displayed. All PFCAs contain a carboxylic acids group and a perfluorinated carbon chain. The compounds differ only in the number of carbon atoms within the fluorinated carbon chain. Thus, we conclude, that all the C8-14-PFCAs belong to the same chemical class and contain not only a common functional group but are highly similar according to their chemical structure.

Table 14: CAS-Numbers and similarity of chemical structures of long chain PFCAs.

Name Abbreviation CAS-No IUPAC Name Chemical structure

PFOA C8-PFCA 335-67-1 Octanoic acid, pentadecafluoro-

CF3(CF2)6-COOH

PFNA C9-PFCA 375-95-1 Nonanoic acid, heptadecafluoro-

CF3(CF2)7-COOH

PFDA C10-PFCA 335-76-2 Decanoic acid, nonadecafluoro-

CF3(CF2)8-COOH

PFUnDA C11-PFCA 2058-94-8 Undecanoic acid, heneicosafluoro-

CF3(CF2)9-COOH

PFDoDA C12-PFCA 307-55-1 Dodecanoic acid, tricosafluoro-

CF3(CF2)10-COOH

PFTrDA C13-PFCA 72629-94-8 Tridecanoic acid, pentacosafluoro-

CF3(CF2)11-COOH

PFTeDA C14-PFCA 376-06-7 Tetradecanoic acid, heptacosafluoro-

CF3(CF2)12-COOH

Dissociation of C8-14-PFCAs and its salts in aqueous media

Under environmental conditions in aqueous media the free perfluorinated carboxylic acids (PFCAs) stay in equilibrium with their conjugate bases, the perfluorinated carboxylates. The fraction of each species depends on the acid dissociation constant (pKa) and the pH of the environmental compartment. Salts of PFCAs, which are sometimes used in laboratory experiments, will be in equilibrium with the corresponding acid in aqueous phases as well. Currently used techniques for analysis and quantification of PFCAs in i.e. environmental samples are not able to distinguish between both of the species. Therefore, reported concentrations always include the acids as well as the bases. If reported concentrations are used for the determination of bioaccumulation factors or for experiments determining the persistency, aqueous phase concentrations include both species. Experimental determination of pKa is difficult for PFCAs, i.e. because of the surface active properties. Calculated values should be taken with care, because for most of the models it is unclear whether PFCAs are within their applicability domain. For assessing the intrinsic properties of the PFCA within this


42

dossier the exact knowledge of the fraction of each species is not required, because both of the species will be available independently from the starting conditions.

Furthermore, due to the uncertainties of pKa values it is not wise to calculate partition coefficients under environmental pH conditions. We would like to mention that there is an ongoing scientific discussion showing that the partitioning of PFCAs in the environment can be described by the properties of the neutral PFCAs only (Webster and Ellis 2011).

Physicochemical properties and partition coefficients of C8-14-PFCAs and some salts

The experimental determination of i.e. partition coefficients is difficult for example because of the surface active properties of the ionic PFCAs. The presence of ionic PFCAs depends on the dissociation of PFCAs in aqueous media. Nevertheless, there are models available, i.e. COSMOtherm calculating partitioning coefficients of neutral PFCAs. COSMOtherm is a quantum chemistry-based method that requires no specific calibration. This calibration would be difficult because of missing measured data of PFCAs. Therefore COSMOtherm is expected to be able to estimate properties for PFASs. Studies have shown that properties estimated with COSMOtherm showed good agreement with the experimental data for a number of per- and polyfluorinated chemicals, i.e. C8-PFCA (Wang et al. 2011; Arp et al. 2006). Again, whether neutral PFCAs are present in aqueous media depends in the dissociation of the acids. Air-water as well as octanol-water partition coefficients are of course different for PFCAs with 8 to 14 carbon atoms but they show a clear increasing, trend with chain length (see Table 15 below, Wang et al., 2011). This can be explained by the increasing molecular volume with each additional CF2-unit. The trend of the fate of PFCAs with chain length is supported by information on sorption of PFCAs on sediment. Sorption increases with increasing chain length (Higgins and Luthy, 2006) also under environmental conditions (Ahrens et al., 2010) (see Table 15).


43

Table 15: Basic substance information and physical chemical properties relevant to justify read across in the PBT assessment.

Abbreviation C8-PFCA C9-PFCA C10-PFCA C11-PFCA C12-PFCA C13-PFCA C14-PFCA

Acronym PFOA APFO NaPFO PFNA PFDA PFUnDA PFDoDA PFTrDA PFTeDA

IUPAC Name Octanoic

acid, pentadecaflu

oro-

ammonium

pentadeca-fluoro-

octanoate

pentadeca

octanoic acid sodium salt

Nonanoic acid,

heptadeca-fluoro-

Decanoic

acid, nonadeca-

fluoro-

Undecanoic

acid, heneicosa-

fluoro-

Dodecanoic

acid, tricosafluoro-

Tridecanoic

acid, pentacosa-

fluoro-

Tetradecanoic

acid, heptacosa-

fluoro-

Chemical Structure

CF3(CF2)6-COOH

CF3(CF2)6-COO-NH4+

CF3(CF2)6-COO-Na+

CF3(CF2)7-COOH

CF3(CF2)8-COOH

CF3(CF2)9-COOH

CF3(CF2)10-COOH

CF3(CF2)11-COOH

CF3(CF2)12-COOH

CAS No 335-67-1 3825-26-1 335-95-5 375-95-1 335-76-2 2058-94-8 307-55-1 72629-94-8 376-06-7 Physico-chemical data

Molecular Weight g/mol

414.09 431.1 464.08 514.08 564.0909 614.0984 664.1059 714.11

2.3 – 2.48 (exp) 2.65 – 2.87 (exp)

3.19 – 3.41 logP 9.363±0.888 at 25°C (calc)

logP 10.093±0.901 at 25 °C (calc)

logP 10.823±0.914 at 25 °C (calc)

Partitioning Coefficient logKow

5.30 (calc., COSMOtherm, Wang et al., 2011)







log KOA 7.23 (calc., COSMOtherm, Wang et al., 2011)







log KAW -1.93 (calc., COSMOtherm, Wang et al., 2011)

-1.58 (calc., COSMOtherm, Wang et al., 2011)






Dissociation constant

2.5 2.8 in 50% aqueous ethanol 1.5 - 2.8

0.52±0.10; (calculated)

0.52±0.10 (calculated)



Partition coefficients log Kd (sediment and overlapping dissolved phase)

0.04 (Ahrens et al., 2010)*

0.6 (Ahrens et al., 2010) *

1.8 (Ahrens et al., 2010) *

3.0 (Ahrens et al., 2010) *

Log Koc (sediment organic

2.06 (Higgins and Luthy, 2006)#

2.39 (Higgins and Luthy, 2006) #

2.76 (Higgins and Luthy, 2006) #

3.3 (Higgins and Luthy, 2006) # 4.8 (Ahrens et


44



carbon-normalized distribution coefficient)

1.09 (Ahrens et al., 2010) *

2.4 (Ahrens et al., 2010) *

3.6 (Ahrens et al., 2010) *

al., 2010) *

Water solubility

9.5 g/L (25° C) 4.14 g/L (22°C)

0.033 mol/L, 14.2 g/L at 2.5 oC (Nielsen 2012)

0.036 mol/L at 8.0 oC at critical micelle concentration (Nielsen 2012)

1.2E-4 g/L; pH 1 at 25°C 9.0E-4 g/L; pH 2 at 25°C 8.5E-3 g/L; pH 3 at 25°C 0.056 g/L; pH 4 at 25°C 0.14 g/L; pH 5 at 25°C 0.16 g/L; pH 6-10 at 25°C (calculated)

2.9E-5 g/L pH 1 at 25°C 2.2E-4 g/L pH 2 at 25°C 2.0E-3 g/L pH 3 at 25°C 0.014 g/L pH 4 at 25°C 0.034 g/L pH 5 at 25°C 0.039 g/L pH 6 at 25°C 0.040 g/L pH 7 at 25°C 0.041 g/L pH 8-10 at 25°C (calculated)

7.3E-6 g/L; pH 1 at 25 °C 5.5E-5 g/L; pH 2 at 25 °C 5.1E-4 g/L; pH 3 at 25 °C 3.5E-3 g/L; pH 4 at 25 °C 8.6E-3 g/L; pH 5 at 25 °C 0.0100 g/L; pH 6-10 at 25 °C (calculated)

1.9E-6 g/L; pH 1 at 25°C 1.4E-5 g/L; pH 2 at 25°C 1.3E-4 g/L; pH 3 at 25°C 9.3E-4 g/L; pH 4 at 25°C 2.2E-3 g/L; pH 5 at 25°C 2.6E-3 g/L; pH 6-10 at 25°C (calculated)

Vapour pressure

4.2 Pa (25 °C) for PFOA extrapolated from measured data 2.3Pa (20 °C) for PFOA extrapolated from measured data 128 Pa (59.3 °C) for PFOA measured

0.0081 Pa at 20 °C, calculated from measured data <0.1 hPa at 20 °C 0.012 Pa at 25 °C 0.0028 Pa at 25 °C (Nielsen 2012)

0.6 to 99.97 kPa (112 to 237.7°C) (calculated)

9.40E-3 Torr at 25°C(calculated)

3.59E-3 Torr at 25°C (calculated)

1.37E-3 Torr at 25 °C (calculated)

Stability

Phototransformation in water DT50

No photodegradation detected under relevant env. conditions

No photodegradation detected under relevant env. conditions

No photodegradation tested under relevant env. conditions 100 % after 12 h by use of persulfate ion (S2O82-) in

No photodegradation tested under relevant env. Conditions 100 % after 12 h by use of persulfate ion

No photodegradation tested under relevant env. Conditions 77% after 12 h by use of persulfate ion (S2O82-) in


45



water (S2O82-) in water

water

Hydrolysis DT50

>97 yr

Direct photolysis

No photo-degradation

indirect photolysis

No photo-degradation (H2O2; synthethic humic water, Fe2O3) estimated half-life > 349 days (Fe2O3)

ready biodegradability screening test

not readily biodegradable (OECD 301 C,F)

not readily biodegradable (OECD 301 B)

not readily biodegradable (OECD 301 F)

not readily biodegradable (OECD 301 C)

not readily biodegradable (OECD 301 C)

Simulation tests

No elimination by metabolic processes, mineralization or adsorption

Biodegradation in soil, sediment

No degradation detected

* pH of the water samples analyzed 7.1-8.3 Temp.: 15.3 – 17.7 °C


46

Table 16 : Information on BCF, BMF and TMF of C9-14 PFCAs relevant to justify read across in the B assessment.

Abbreviation C9-PFCA C10-PFCA C11-PFCA C12-PFCA C13-PFCA C14-PFCA

Acronym PFNA PFDA PFUnDA PFDoDA PFTrDA PFTeDA IUPAC Name Nonanoic acid,

heptadecafluoro- Decanoic acid, nonadecafluoro-

Undecanoic acid, heneicosafluoro-

Dodecanoic acid, tricosafluoro-

Tridecanoic acid, pentacosafluoro-

Tetradecanoic acid, heptacosafluoro-

Chemical Structure

CF3(CF2)7-COOH CF3(CF2)8-COOH CF3(CF2)9-COOH

CF3(CF2)10-COOH

CF3(CF2)11-COOH CF3(CF2)12-COOH

CAS No 375-95-1 335-76-2 2058-94-8 307-55-1 72629-94-8 376-06-7 Physico-chemical data

Molecular Weight g/mol

464.08 514.08 564.0909 614.0984 664.1059 714.11

2.3 – 2.48 (exp) 2.65 – 2.87 (exp) logP 9.363±0.888 at 25°C (calc)

logP 10.093±0.901 at 25 °C (calc)

logP 10.823±0.914 at 25 °C (calc)

Partitioning Coefficient log KOW







log KOA 7.50 (calc., COSMOtherm, Wang et al., 2011)






log KAW -1.58 (calc., COSMOtherm, Wang et al., 2011)






BCF

Lumbriculus variegatus / sediment

0.64 ± 0.05 0.06 ± 0.04

Rainbow trout (carcass)

450 ± 6 62 2700 ± 400 18000 ± 2700 23000 ± 5300

Rainbow trout (blood)

2700±350 11000 ± 1400 40000 ± 4500 30000 ± 4200

Rainbow trout (liver)

1100 ± 180 4900 ± 770 18000 ± 2900 30000 ± 6000

Carp (whole) 2300 - 3700 10000-16000 16000 - 17000 Juvenile rainbow trout (carcass) calculated*

4044-6326 5761-8210 10388-19294


47

Juvenile rainbow trout (liver) calculated*

5245-6071 7926-8576 17835-19294

BAF Chup(Leuciscus cephalus) plasma

1000000 5011872

Chup(Leuciscus cephalus) liver

100000 501187

Chup(Leuciscus cephalus) gills

79433 501187

Chup(Leuciscus cephalus) gonads

63096 316228

Chup(Leuciscus cephalus) muscle

19953 100000

BMF 0.13-111 0.21-87 0.21 - 353 0.1 – 156 0.35 – 9 0.33 – 8.5 TMF 1.9-7.0 1.5-20 0.75 – 31.2 0.6 – 3.76 1.4 – 2.45 0.23 – 2.6 * Calculated BCFs based on BMF values for C11-PFCA (0.28 ± 0.04), C12-PFCA (0.43 ± 0.062) and C14-PFCA (1.0 ± 0.25).

# pH of sediment analysed: 5.7 to 7.6


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7 References

3M Co. 1978. Biodegradation (ABS/LAS Shake Culture Test). St. Paul, MN. Project number

9970612613. U.S. Environmental Protection Agency Administrative Record 226-0489.

3M Co. 1979. Photolysis study using simulated sunlight. FC-143 Photolysis study using

simualted sunlight U.S. Environmental Protection Agency Administrative Record 226-0490.

3M Co. 1980. Environmental Laboratory Acute Toxicity Testing: FC-12. U.S. Environmental

Protection Agency Administrative Record 226-0504.

3M Co. 1985. Ready Biodegradation of FX-1001 (BOD/COD). Lab Request No. C1006.

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